UNEP-WCMC technical report
Review of species selected on the basis of the Analysis of the European Union annual reports to CITES 2017
(Version edited for public release)
Review of species selected on the basis of the Analysis of the European Union annual reports to CITES 2017.
Prepared for The European Commission, Directorate General Environment, Directorate F - Global Sustainable Development, Unit F3 - Multilateral Environmental Cooperation, Brussels, Belgium.
Published June 2019
Copyright European Commission 2019
Citation UNEP-WCMC. 2019. Review of species selected on the basis of the Analysis of the European Union annual reports to CITES 2017. UNEP-WCMC, Cambridge.
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Contents
Introduction ...... 2
Lama guanicoe II/B ...... 3
Crocodylus niloticus II/B ...... 23
Sphyrna lewini II/B ...... 36
Appendix ...... 47
i
Introduction
Introduction
This report presents a review of three species selected on the basis of the Analysis of the European Union annual reports to CITES 2017.
On the basis of the 2017 Analysis, 123 taxa were identified as having noteworthy trends in EU imports of wild or ranched specimens, from which four Appendix II/Annex B species were suggested for review:
Lama guanicoe (Guanaco) – Chile Caiman crocodilus yacare (Yacare caiman) – Brazil Crocodylus niloticus (Nile crocodile) – Malawi Sphyrna lewini (Scalloped hammerhead) – Kenya
These four taxa/country combinations were selected by the SRG for review, along with Lama guanicoe from Argentina. The assessment for Caiman crocodilus yacare from Brazil will be presented at the next SRG meeting, alongside reviews for Bolivia and Paraguay, which were separately selected for in-depth review on the basis of long-standing opinions from those range States.
2 Lama guanicoe
ARTIODACTYLA: CAMELIDAE
Lama guanicoe II/B
SYNONYMS: Lama glama guanicoe Müller, 1776
COMMON NAMES: Guanaco (EN, ES, FR)
RANGE STATES: Argentina, Plurinational State of Bolivia, Chile, Falkland Islands (Malvinas) (Introduced), Paraguay, Peru
UNDER REVIEW: Argentina, Chile
EU DECISIONS: Current no opinion i) for wild-sourced specimens from Chile other than the population of Tierra del Fuego formed on 07/11/2016, replacing a previous no opinion with the same annotation formed on 15/03/2005
Current positive opinion for wild-sourced specimens from Chile for the population on the Guanaco Conservation and Management Program, Region XII, Magallanes, Tierra del Fuego, formed on 08/12/2014, replacing a previous positive opinion for animals and products from the same population formed on 09/10/2003
Previous Article 4.6(b) import suspension for wild-sourced specimens from Chile formed on 03/02/2001 and last confirmed on 01/03/2003
Previous Article 4.6(b) import suspension for wild specimens from Argentina (except (i) specimens that form part of the registered stock in Argentina, provided that permits are confirmed by the Secretariat before being accepted by the Member State of import; (ii) products obtained from the shearing of live animals approved management programme, appropriately marked and registered; (iii) non-commercial exports of limited quantities of wool for industrial testing, up to 500 kg annually) formed on 19/09/1999 and last confirmed on 03/09/2008
IUCN: Least Concern
Taxonomic note
The species was formerly listed in CITES as Lama glama guanicoe. Historically, four subspecies were recognised based on skull measurements, coat colouration, body size and distribution (Franklin, 2011). There are currently only two subspecies which are recognised, based on recent molecular studies (González et al., 2006): Lama guanicoe cacsilensis (Lönnberg, 1913) and Lama guanicoe guanicoe (Müller, 1776). The current CITES Standard Reference for Camelidae recognises L. guanicoe as a separate species to L. glama, and includes L. cacsilensis as a synonym (Wilson and Reeder, 1993). Wilson and Mittermeier (2011) note that a significant biogeographic revision of the two current subspecies is required.
3 Lama guanicoe
Trade patterns
Lama guanicoe was listed in CITES Appendix II on 12/08/1978 and in Annex B of the European Union (EU) Wildlife Trade Regulations on 01/06/1997; in both cases the listing was under its former accepted name, Lama glama guanicoe, which is now considered a synonym.
Argentina
Argentina has submitted all annual reports for the period 2008-2017 with the exception of 2014. Argentina has never published export quotas for L. guanicoe.
According to the CITES Trade Database, direct trade to the EU-28 2008-2017 predominantly consisted of hair and fibres for commercial purposes: with 4092 kg reported by Argentina and 4972 kg reported by importers (Table 1). Over two thirds of trade was wild-sourced and the remainder was captive-bred.
Direct exports to the rest of the world 2008-2017 primarily comprised hair reported 2008-2011 and in 2017, the vast majority of which was wild-sourced for commercial purposes. Other exports to the rest of the world since 2011 were limited to low levels of wild-sourced garments and cloth for commercial purposes and specimens for scientific purposes (Table 1). Indirect trade to the EU-28 originating in Argentina 2008-2017 primarily constituted low levels of wild-sourced garments for commercial purposes (Table 2), most of which were re-exported via Switzerland.
4 Lama guanicoe
Table 1: Direct exports of Lama guanicoe from Argentina to the EU-28 (EU) and the rest of the world (RoW), 2008-2017. Quantities have been rounded to whole numbers where applicable. Due to low levels of trade (<5 units), trade in plates, small leather products and skin pieces, as well as trade reported as purpose ‘P’ has been excluded from the table. Importer Term Unit Purpose Source Reported by 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 Total EU cloth kg T C Importer Exporter 48 48 W Importer 123 123 Exporter 163 163 fibres kg T C Importer 31 31 Exporter W Importer 663 298 221 139 1 1322 Exporter 221 221 garments - T W Importer Exporter 1 32 33 hair kg T C Importer 543 98 360 179 1180 Exporter 456 22 569 212 31 1290 F Importer 1 1 Exporter W Importer 223 775 31 1400 8 2437 Exporter 151 785 123 1382 139 1 2581 skins kg T C Importer Exporter 88 88 W Importer Exporter 72 72 - T W Importer 1 1 Exporter specimens - S W Importer Exporter 29 29 RoW cloth kg T W Importer Exporter 202 202 m T W Importer 10 10 Exporter garments - T W Importer 8 8 Exporter 6 6 hair kg T C Importer 47 47 Exporter 47 47 W Importer 550 253 11 210 151 1175
5 Lama guanicoe
Importer Term Unit Purpose Source Reported by 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 Total Exporter 300 254 215 769 skins kg T C Importer Exporter 49 49 W Importer Exporter 201 201 specimens - S W Importer Exporter 40 40 T W Importer Exporter 42 42 teeth - S W Importer Exporter 179 179 Source: CITES Trade Database, UNEP-WCMC, Cambridge, UK, downloaded on 03/06/2019.
Table 2: Indirect exports of Lama guanicoe originating in Argentina to the EU-28, 2008-2017. All trade was for commercial purposes. Quantities have been rounded to whole numbers, where applicable.
Term Unit Source Reported by 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 Total cloth m² W Importer 3 3 Exporter - W Importer 6 6 Exporter 1 1 derivatives - W Importer Exporter 4 4 garments - C Importer 1 1 Exporter W Importer 3 12 2 2 10 17 2 9 57 Exporter 12 13 2 8 22 6 8 71 leather products - W Importer 12 7 19 (small) Exporter Source: CITES Trade Database, UNEP-WCMC, Cambridge, UK, downloaded on 03/06/2019.
6 Lama guanicoe
Chile
Chile has submitted all annual reports for the period 2008-2017. Chile has never published quotas for L. guanicoe. According to the CITES Trade Database, direct exports to the EU-28 2008-2017 predominantly consisted of wild-sourced meat for commercial purposes to the Netherlands: ~333 000 kg as reported by Chile and ~292 000 kg as reported by the Netherlands (Table 3). Exports of meat in 2017 increased more than five times compared to 2016, and were nearly triple levels reported in 2014, the second highest year of trade 2008- 2017. Direct exports of L. guanicoe to the rest of the world 2008-2017 principally comprised 500 wild-sourced skins for commercial purposes exported to Uruguay in 2008, as reported by Chile; Uruguay reported this import in 2009.
Indirect trade in L. guanicoe to the EU-28, originating in Chile 2008-2017 consisted of three captive-sourced garments re-exported via Switzerland to France in 2010, for commercial purposes.
7 Lama guanicoe
Table 3: Direct exports of Lama guanicoe from Chile to the EU-28 (EU) and the rest of the world (RoW), 2008-2017. Importer Term Unit Purpose Source Reported by 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 Total
EU garments - T W Importer Exporter 10 10 meat kg T W Importer 12400 22000 32850 45216 21750 28000 129867 292083 Exporter 44 56400 50000 33500 50 21750 28000 143000 332744 skins - T W Importer 75 10 10 95 Exporter 77 10 87 specimens - S C Importer 800 800 Exporter W Importer 500 500 Exporter RoW skins - T W Importer 500 500 Exporter 500 500 Source: CITES Trade Database, UNEP-WCMC, Cambridge, UK, downloaded on 03/06/2019.
8 Lama guanicoe
Conservation status
Lama guanicoe is the largest, most widespread and most abundant native camelid in South America (Franklin, 1982; Castillo et al., 2018; Ortega and Franklin, 1995). Measurements of the species vary across its distribution, but it reaches its maximum size (approx. 140 kg) in southern Chile and its smallest size (approx. 90 kg) in Northern Peru (Wilson and Mittermeier, 2011). L. guanicoe is a non-specialised grazer, which forages on a wide range of species (Burgi et al., 2012) and may browse on shrubs and trees during winter (Ortega and Franklin, 1995). It occurs in three predominant types of social group: family groups, solitary males and male groups (Eisenberg and Redford, 2000). Group sizes were reported to be variable; between three to 60 individuals for male groups, averaging 25 individuals, and mixed groups numbering up to 500 individuals (60 on average) (Wilson and Mittermeier, 2011). Females can breed from their second year, producing a single offspring after 11.5 months of gestation and longevity is c. 28 years (Sarno et al., 1999; Wilson and Mittermeier, 2011).
Both sedentary and migratory populations exist in South America, with migration thought to be driven by snow depth and winter forage supply (Ortega and Franklin, 1995). Migratory home ranges can reach up to 900 km² in the Andes (Baldi et al., 2016) and 40 km² in Tierra del Fuego (Moraga et al., 2014), compared to 2-9 km² (average 4 km2) in sedentary individuals located in Figure 1: Range of Lama guanicoe. Source: Baldi et al., 2016. Chubut Province in Argentina (Marino and Baldi, 2008). L. guanicoe typically lives in highly seasonal environments, including montane habitats up to 4500 m a.s.l., although it also occurs at sea level at higher latitudes (Franklin, 2011). The species inhabits four of the 10 principal habitats in the continent: desert and xeric shrublands, montane and lowland grasslands, savannahs and shrublands, and temperate forest (which includes regions such as the southern Pampa) (Wilson and Mittermeier, 2011). Southern Chile’s arid habitats are home to vegas, or meadows, that are the preferred foraging grounds of female guanacos (Wilson and Mittermeier, 2011). These habitats provide the additional benefit of protection against predators; namely pumas (Puma concolor), which do not typically inhabit these sites (Bank et al., 2003).
L. guanicoe ranges discontinuously from north-western Peru to the southern tip of the continent, in Tierra del Fuego (8° to 55°S) (Fig. 11; González et al., 2006; Baldi et al., 2009; Franklin and Fanklin, 1982; Baldi et al., 2016), with a total estimated range of c. one million km² (Baldi et al., 2016). It is native to Argentina, Bolivia, Chile and Peru (Redford and Eisenberg, 1992), as well as Paraguay: a small population has been reported to occur in the western Paraguayan Chaco (Yahnke et al., 1998). L. guanicoe was also introduced to Staats Island in the Falkland Islands in the 1930s, where a population of around 400 individuals persists (Franklin and Grigione, 2005). The Patagonian grasslands and shrublands still harbour most of the world’s L. guanicoe population (Baldi et al., 2009), with the vast majority of the global population distributed in Argentina
1 Disclaimer: The boundaries and names shown and the designations used on maps do not imply official endorsement or acceptance by UN Environment or contributory organisations.
9 Lama guanicoe
(González and Acebes, 2016; Wilson and Mittermeier, 2011). The Peruvian and northern Chilean populations were reported to be L. guanicoe cacsilensis and the Bolivian, Argentinian, central and southern Chilean populations were attributed to L. guanicoe guanicoe (Marín et al., 2008; Franklin, 2011). A hybridised population is present between the two subspecies’ distribution ranges (Marín et al., 2013), between approximately 26° and 32°S (Castillo et al., 2018). L. guanicoe’s current range is considered to be less than 40% of its original range (Franklin, 1982 in: Ceballos and Ehrlich, 2002; Wilson and Mittermeier, 2011), and remaining populations have been reported to be often isolated and fragmented (González et al., 2006; Marín et al., 2013). Castillo et al. (2018) further predict that the species could lose an additional 21% of its current geographic distribution by 2070. The IUCN reported that L. guanicoe may become extinct in Bolivia, Paraguay and Peru, which together share less than 1% of the continent’s total population (González and Acebes, 2016; Baldi et al., 2016).
South America’s L. guanicoe population is estimated to be 3-7% of the population present before European colonisation in the 17th century (30-50 million individuals), primarily due to uncontrolled hunting for hides and meat and the introduction of domesticated sheep (Raedeke, 1979; Marín et al., 2013). The most recent IUCN Red List assessment of L. guanicoe estimated the global population to be between 1.5 and 2.2 million individuals (with an estimated 1-1.5 million adults) “based on life-tables of Raedeke 1979 and Fritz and Franklin 1994”, with a current population trend of ‘increasing’ (Baldi et al., 2016). Whilst multiple actions focussed on reversing the species’ critical situation in the 1960s and 1970s have resulted in the recovery of some populations, the methodology used to calculate the recent population estimate was noted to be based on a multitude of studies which used different sampling and analysis methods (González and Acebes, 2016). Travaini et al. (2015) noted that L. guanicoe densities are highly variable across regional scales, thus requiring large-scale surveys for accurate regional population estimates. As it stands, population estimates across L. guanicoe’s entire geographic range are extrapolated from many small-scale surveys, resulting in potential inaccuracies. The figure is close to three times higher than the 535 750 – 589 750 individuals calculated by an IUCN Red List assessment of the population in 2008 (Baldi et al., 2008). It was reported that this increase may be partly attributed to the use of new information and novel ways of estimating population sizes (e.g. Schroeder et al., 2014; Travaini et al., 2015), alongside the net growth of populations in Chilean Patagonia (Zubillaga et al., 2014a) and inclusion of recent surveys from some large areas previously unsurveyed (Baldi et al., 2016). The IUCN cautioned that, whilst the overall current population trend was considered ‘increasing’, populations were “small and in real decline or at best tenuously stable” outside the far southern cone of South America (Baldi et al., 2016).
L. guanicoe was categorised as Least Concern in a 2016 IUCN assessment, on the basis of its extensive distribution, presumed population size and the occurrence of large populations within numerous protected areas spread across its range (Baldi et al., 2016). The assessment noted that the status of populations within the three countries where L. guanicoe was currently classified as ‘endangered’ (Paraguay, Bolivia and Peru) was “of grave concern” (Baldi et al., 2016).
Hunting has been the predominant threat to the species historically, negatively affecting populations regardless of their size and density (Travaini et al., 2015). Current threats include hunting, poaching and disturbance by landowners (Burgi et al., 2012; Baldi et al., 2009), as well as habitat degradation and fragmentation caused by sheep overgrazing and infrastructural development (Gable and Durham, 2016; Rey et al., 2012; Burgi et al., 2012) and extractive industries (Baldi et al., 2016). Entanglement in fences was also reported as a threat (Rey et al., 2012). Population declines have also been attributed to management plans which were not built to accommodate the needs of both ranchers and conservation (Gable and Durham, 2016).
L. guanicoe was considered to have a high economic value as a commercial species, for hair, meat and skin products (Baldi et al., 2009; Zubillaga et al., 2014a). Both males and females have woolly coats prized for the value of their soft fibres (Mueller et al., 2015). In 2015, L. guanicoe fibre production from Argentina was estimated at over 2000 kg, with recommendations to increase production based on recent population increases (Mueller et al., 2015). The fibre has been valued at USD 100-200 per kg for unclean wool, and
10 Lama guanicoe
USD 400 per kg for cleaned and de-haired wool (Wilson and Mittermeier, 2011). The average quantity of wool from a carcass was reported to be 0.32-0.35 kg of wool (Wilson and Mittermeier, 2011). Current conservation measures were considered to be ad hoc rather than holistic, focussing on the recurring “emergencies” presented by intensive local poaching, rather than considering the full range of threats across the species’ vast range (Baldi et al., 2016; González and Acebes, 2016). Programmes that allow the maintenance of abundant and functional populations of L. guanicoe were considered vital for the conservation of the species, and it has been suggested that effective conservation measures must be implemented at the local scale, which would take the variation in population densities into consideration (Marín et al., 2013; Baldi et al., 2016). Conservation efforts focussed on the recovery of small and fragmented populations and their connectivity was also recommended, alongside the sustainable use of recovered and abundant populations (González and Acebes, 2016).
Argentina
Argentina holds more than 80% of the global population of Lama guanicoe, estimated between 1.23 and 1.89 million individuals (Franklin, 2011; González and Acebes, 2016) (Table 4). The species is distributed across the near-totality of the Argentine Patagonia, but it is fragmented into small and relatively isolated populations (Burgi et al., 2012). L. guanicoe populations are more fragmented in the northern Patagonian provinces (Chubut, Mendoza, Neugén and Río Negro), than in the south (Santa Cruz and Tierra del Fuego provinces) (González and Acebes, 2016). The L. guanicoe populations in central and northern Argentina are restricted to the pre-Andean and Andean mountain ranges reaching up to the Bolivian border (Baldi et al., 2016). In 2007, just over 100 individuals were successfully reintroduced in the Quebrada del Condorito National Park in Córdoba, central Argentina (Barri and Cufré, 2014).
Table 4: Population estimates for Lama guanicoe in Argentinean Provinces, adapted from González and Acebes, 2016. Province Abundance (number of individuals) Source Catamarca 1829 – 2253 Baigún et al., 2008 Chubut 300 000 – 560 166 Gavuzzo et al., 2015, Baldi pers. comm. Córdoba 36 Barri and Cufré, 2014 Jujuy 462 – 696 Baigún et al., 2008 La Pampa 151 – 180 Sosa and Sarasola, 2005 La Rioja 1711 – 1915 Baigún et al., 2008 Mendoza 13 549 – 25 951 Schroeder et al., 2014 Neuquén 30 000 – 35 000 Funes pers. comm. Rio Negro 120 000 – 160 000 Funes pers. comm. Salta 731 – 750 Baigún et al., 2008 San Juan 13 179 – 14 834 Baigún et al., 2008 San Luis 6 – 18 Baigún et al., 2008 Santa Cruz 727 800 – 1 066 600 Travaini et al., 2015 Tierra del Fuego 16 307 – 21 868 Schiavini, pers. comm. Total 1 225 761 – 1 890 267
L. guanicoe is thought to occupy 60% of its historical distribution in Argentina (Franklin et al., 1997). Their historical decline has been attributed to competition with domestic sheep (Ovis aries) (Raedeke, 1979; Baldi et al., 2001) which were introduced by European settlers; the sheep population reached 22 million within 50 years of their introduction, whilst L. guanicoe decreased to fewer than a million individuals by the end of the 1900s (Baldi et al., 2009). L. guanicoe densities were reported to be negatively correlated with domestic sheep numbers (Schroeder et al., 2013). The two species’ diets were reported to overlap significantly in summer when food resources are scarcer than in spring (Gable and Durham, 2016).
11 Lama guanicoe
The overall current population trend of L. guanicoe in Argentina has been increasing (Travaini et al., 2015; Baldi et al., 2016), which was partly attributed to advances in population surveying methods (Travaini et al., 2015) as well as increased numbers in managed reserves such as the Reserva San Pablo de Valdés (Burgi et al., 2012). The increase in L. guanicoe density in this reserve may have been due to the removal of sheep, as well as a permanent warden onsite who prevented poaching and disturbance of the population (Burgi et al., 2012). In a similar protected area featuring a permanent warden, Reserva Cabo dos Bahías, L. guanicoe density reached 34-44 individuals per km2, an order of magnitude higher than the average density in the rest of the Chubut region (2-9 individuals per km2) (Baldi et al., 2001). Across Argentine Patagonia, L. guanicoe densities were reported to be fewer than five individuals per km2 (Puig et al., 2003; Baldi et al., 2009). In a 2004-2005 road survey of Santa Cruz province, southern Patagonia, density estimates were found to range from 1.1 to 7.7 individuals per km2 (mean value 4.79/km2), with large spatial variations in density (Travaini et al., 2015). In the Chubut, Neuquén and Rio Negro Provinces of Patagonia, densities were found to be lower than average, with fewer than two individuals per km2 (Baldi et al., 2009). The remainder of the country has an average density below one individual per km2, and the populations remain highly fragmented (Baigún et al., 2008a). There was reported to be a need for more up-to-date and accurate estimates of the Argentine population, as differences in survey methods and effort across such a large area have led to uncertainty of the actual numbers (Travaini et al., 2015; González and Acebes, 2016).
At the national level, L. guanicoe was considered ‘least concern’ in 2012, and had been listed as ‘near threatened’ in 1997 and 2000 (Ojeda et al., 2012); however, relict populations were considered under threat due to being highly fragmented (Wilson and Mittermeier, 2011).
The main threats to L. guanicoe in Argentina are grazing competition with domestic sheep (Ovis aries), recreational hunting and poaching, habitat degradation and infrastructure development (Burgi et al., 2012; Rey et al., 2012; Travaini et al., 2015; Baldi et al., 2016). L. guanicoe was reported to be mainly hunted by poachers, who are usually urban residents using vehicles along roads (Rivas et al., 2015). This type of hunting method requires conservation interventions such as closing certain access routes (e.g. unused oil trails) as well as increasing ranger patrols (Rivas et al., 2015). L. guanicoe was also reported to be threatened by increased infrastructure, which has caused entrapment and road kills close to roads, and has repercussions on limiting population movements and local isolation (Rey et al., 2012). L. guanicoe populations in the Auca Mahuida oil field area in northern Patagonia have declined by ~90% from 1986 to 2004, as road density increased from 0.14 to 1.84 km/km² (Radovani, 2004 in: Baldi et al., 2009).
Lama guanicoe pelts had historically been exported by Argentina. An estimated 233 610 pelts were exported between 1976-1979 (Iriarte and Jaksic, 1986). In 1979, the value of L. guanicoe pelt exports since 1972 amounted to USD 3.6 million, and permits were reportedly issued in Chubut region to hunt ~118 000 animals between 1984 and 1994, based on population estimates provided by landowners (Baldi et al., 2009). The decrease in hunting and exports of pelts since then has been attributed to the implementation of international and national regulations; L. guanicoe was listed in CITES Appendix II in 1992, and pelt and fibre exports were prohibited under national Resolution 220/98 (Government of Argentina, 1998) until the implementation of a sustainable management plan (Baldi et al., 2009). The same resolution (220/98) (Government of Argentina, 1998) also prohibited the exportation, commercialisation and interprovincial transit of live individuals, products and by-products of L. guanicoe until a management plan for sustainable use of the species was adopted. It provided exemptions for stocks of L. guanicoe hides registered prior to February 1994, products and by-products that were the issue of live-shearing, and any products exported for non-commercial purposes, up to 500 kg per company.
The National Management Plan for Guanacos, proposed in 1998 and implemented in 2006 (Baldi et al., 2006), included regulations on hunting and sustainable use of L. guanicoe, comprising the harvesting of wool by local farmers as an alternative source of revenue (von Thüngen and Lanari, 2010). It was coordinated by the Federal Wildlife Agency (Dirección de Fauna Silvestre), and implemented multiple provincial acts as well as the Federal Wildlife Conservation Law to create a legal basis for protecting and sustainably using the species. The Federal
12 Lama guanicoe
Wildlife Conservation Law (Government of Argentina, 1981) allows for the use of wildlife temporarily or permanently inhabiting owned land, whilst remaining in accordance with other national and provincial laws and regulations.
Permits are required to hunt wildlife, including L. guanicoe, and there are imprisonment penalties for infringing upon the hunting regulations (up to three years) as well as fines up to 50 million pesos (Government of Argentina, 1981). L. guanicoe is one of the species protected by Provincial Law 101 (Government of Argentina, 1993), which prohibits hunting in the region of Tierra del Fuego, Antarctica and the Southern Atlantic Islands. In late 2018, an Argentinean newspaper reported that the Tierra del Fuego governor had vetoed modifications to the protections that Law 101 afforded to L. guanicoe, thus ensuring that hunting continued to be prohibited (Infobae, 2018).
Shortcomings of the National Management Plan have been noted by researchers and non-governmental workers, who have pointed to its inefficiency and unsuccessful implementation, due to a lack of understanding of local farmers’ needs, who were most directly affected by the legislation (Gable and Durham, 2016; von Thüngen and Lanari, 2010). The management plan included provisions on the utilisation of L. guanicoe for meat and wool (Baldi et al., 2006). Around 35% of the high density Argentine populations that have been identified are under programs that call for live shearing for sustainable use for their wool (Wilson and Mittermeier, 2011). However, under this Management Plan farmers require special permits for L. guanicoe wool harvesting, which are costly in comparison to sheep farming, which requires no special permits. This financial hurdle has prevented many farmers from transitioning to L. guanicoe management (von Thüngen and Lanari, 2010). Furthermore, this group of stakeholders has been reported to view L. guanicoe as a pest due to reports indicating that L. guanicoe and O. aries compete for forage (von Thüngen and Lanari, 2010). Live- shearing of captive and wild L. guanicoe was popular between 2004-2008, where over ten thousand L. guanicoe were shorn, and has since decreased due to variations in the global market price for crude fibre (Lichtenstein, 2013; Baldi et al., 2016). Live shearing of wild animals, that are subsequently released, had been suggested as a more sustainable method of making economic profits from L. guanicoe, compared with the shearing of hunted animals or shearing in closed-management systems, where juvenile L. guanicoe less than a year old (‘chulengos’; Travaini et al., 2015) are taken from the wild to stock farm systems (Baldi et al., 2009).
The ‘National Plan for the Sustainable Management of Guanaco‘ is in the process of being updated and a ‘Project for the sustainable use of wild guanacos’ recently involved a pilot test on sustainable commercial exploitation in the province of Santa Cruz (Argentinian Committee of IUCN, 2019). This included the production and export of >19 tonnes of meat to the EU-28 in 2018 (Argentinian Committee of IUCN, 2019). Resolution 766/2017 of the Secretariat of Environment and Sustainable Development (SAyDS) authorized the export, interprovincial transit and domestic marketing of products and by-products of up to 6000 L. guanicoe individuals harvested during 2018 within the framework of this project (Ministry of Environment and Sustainable Development, 2017). This represented a large increase over the 200 individuals authorized for harvest in October 2017 under Resolution 711/2017 (Argentinian Committee of IUCN, 2019).
Concerns have been raised by researchers, civil society organisations and the IUCN/SSC Camelidae Specialist Group (GECS) regarding the current management practices for L. guanicoe, including the lack of civil society participation in development of the national management plan2 and in relation to the pilot test and commercialisation of meat products (Argentinian Committee of IUCN, 2019). It was noted that sustainable use initiatives of wild guanacos needed to be developed with meticulousness and care, evaluating environmental, social and economic impacts, seeking citizen participation and involving updated and validated guanaco population censuses (Argentinian Committee of IUCN, 2019). Likewise, Travaini et al. (2015) recommended that the sustainability of potential L. guanicoe exploitation should be based on sound population estimates, as well as a well-designed monitoring protocol, prior to commencing any extractive activity. The Guanacos de
2 An online public consultation was later launched.
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Santa Cruz Management Plan (PMGSC) was also criticised for asserting that L. guanicoe has a fundamental role in desertification, promoting extractive management measures (Argentinian Committee of IUCN, 2019).
Less than 1% (as of 2015, 0.7%) of the Patagonian steppe is estimated to be under effective protection (Rivas et al., 2015). Protected areas in the Patagonian steppe have been estimated to cover 10% of the L. guanicoe population (Baldi et al., 2016); however, concerns have been raised regarding the effectiveness of protected areas as a result of grazing of competing livestock, a lack of wardens, and a high level of poaching (Wilson and Mittermeier, 2011).
Chile
Lama guanicoe is distributed from Chile’s northern border with Peru down to Navarino Island in the extreme south (González et al., 2013). L. g. cacsilensis and L. g. guanicoe both occur in the country (Franklin, 2011; González et al., 2013; Marín et al., 2013): small and fragmented populations of L. g. cacsilensis occur in the north (Marín et al., 2008; González et al., 2013), whereas L. g. guanicoe is primarily concentrated in the Magallanes and Aysén regions of the far south (González et al., 2013), where the largest populations persist (González and Acebes, 2016). A hybrid zone between the two subspecies occurs in central Chile (González et al., 2013; Marín et al., 2013).
Chile holds 14-18% of South America’s guanaco population (Baldi et al., 2016); an estimated 270 000 to 299 000 individuals (Table 5; González and Acebes, 2016), though these figures were considered ‘speculative’ with numbers sourced from ‘scattered information rather than planned surveys’ (Baldi et al., 2016). Population estimates for the comunas of the Magallanes region in 2018 are outlined in Table 6.
Table 5: Population estimates for Lama guanicoe in Chilean Regions, adapted from González and Acebes, 2016. Abundance (number Region Area Source of individuals) Antofagasta Whole region 244 – 2986 González, unpublished data Araucanía Alto Biobío 40 Cunazza, 1991 Arica y Parinacota Pre-cordillera 1231 Sielfeld, 2004 Atacama Llanos de Challe 900 CONAF, unpublished data Nevado Tres Cruces 68 González, unpublished data Pan de Azúcar 24 – 66 CONAF, unpublished data Pascua 195 González, unpublished data Vallenar-Copiapó 62 González, unpublished data Other 100 – 300 González, unpublished data Aysén Valle Chacabuco 2800 – 3200 Saucedo, unpublished data Other 50 – 150 González, unpublished data Coquimbo Choros 320 González, unpublished data Pelambres 1321 González, unpublished data Other 100 – 300 González, unpublished data Magallanes Isla Navarino 50 – 100 González, unpublished data San Gregorio 63 850 Soto, unpublished data Tierra del Fuego 183 000 – 197 000 Soto, 2010 Torres del Paine 12 801 – 21 966 Iranzo, unpublished data O’Higgins Cipreses 300 – 400 CONAF, unpublished data Tinguiririca 100 – 200 Pavez, unpublished data RM Cruz de Piedra 85 – 200 González, unpublished data Tarapacá Pre-cordillera 104 – 2142 González, unpublished data Valparaíso Alicahue 2267 González, unpublished data Total 270 012 – 299 264
Table 6: Population estimates for Lama guanicoe from Magallanes region, 2018. Source: Soto Volkart (2019)
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Comuna Number of Individuals Torres del Paine 14 609 San Gregorio 86 551 Primavera 18 149 Porvenir 57 544 Timaukel 77 065
L. guanicoe historically ranged over most of the country (Franklin et al., 1997), but the species’ range in Chile was estimated to have been reduced by 75% from its historical distribution (Franklin et al., 1997). Populations in the far south of Chile have experienced significant growth in recent years (Zubillaga et al., 2014b; Iranzo et al., 2018). In Magallanes, stable populations of L. guanicoe were reported to occur in four provinces of the region (CONAMA, 2009), with the greatest abundance in Tierra del Fuego (see further details below) (CONAMA, 2009; González and Acebes, 2016). The L. guanicoe population located in Torres del Paine National Park was described as the second most important in Chile after Tierra del Fuego by Iranzo et al. (2018). The population within the National Park increased from less than 100 individuals in 1975 (Franklin and Fanklin, 1982 in: Iranzo et al., 2018) to around 4200 in 2010, and was reported to have expanded outside of the park into areas where the species had been absent for ‘many decades’ (Iranzo et al., 2018). Densities recorded 2009-2011 ranged from 16-37 individuals per km2 within the National Park to 8-15 ind./km2 outside (Iranzo et al., 2018).
However, while some high-density populations occur in the south of the country, the remainder of the Chilean populations have been described as ‘small and widely scattered’ (Baldi et al., 2016). The Chilean population of L. g. cacsilensis in the north was reported to live in small, fragmented groups (Marín et al., 2008), with reports of population declines of this subspecies in Azúcar National Park (Atacama Region) (Morales and Grimberg, CONAF Atacama, unpublished data in: CONAMA, 2009) and local extirpations in coastal areas around Antofagasta and south of Iquique (Amado pers. comm. in: CONAMA, 2009).
In 2011, the Ministry of the Environment classified L. guanicoe as ‘vulnerable’ in northern and central Chile (between the Arica and Parinacota region and the Los Lagos region) and ‘least concern’ in southernmost Chile (from the Aysén region to the Magallanes region) (Ministerio del Medio Ambiente, 2012). The species’ classification as ‘vulnerable’ was based on an estimated population decline of 10% (inferred from data collected in the Pan de Azúcar National Park), and on the assumption that none of the subpopulations contained over 1000 adults (González, unpublished data in: Ministerio del Medio Ambiente, 2012). The L. guanicoe subpopulations classified as ‘least concern’ were categorised as such due to the species’ general abundance in these regions, with stable population trends (Ministerio del Medio Ambiente, 2012). According to Regulation of the Hunting Law (DS No.5 of 1998, MINAGRI), L. guanicoe was previously classified as ‘endangered’ in the Arica and Parinacota and Los Lagos regions, and ‘vulnerable’ in the Aysén and Magallanes regions (Government of Chile, 1998).
The major threats to L. guanicoe in Chile were reported to be recreational hunting and poaching (CONAF, 2013; Baldi et al., 2016), predation and disturbance by feral dogs in protected areas (CONAF, 2013; Soto, 2014), and the fragmentation, destruction and modification of the species’ habitat (Ministerio del Medio Ambiente, 2012; CONAF, 2013), with extractive industries and construction of large-scale road networks contributing towards the latter (Valladares Faúndez, 2012; CONAF, 2013).
In addition to the primary threats faced by L. guanicoe, the species was reported to be susceptible to diseases from domestic animals; in northern Chile, L. g. cacsilensis populations have been reportedly affected by scabies, causing high mortality in low density populations (Castillo et al., 2008; Valladares Faúndez, 2012; Baldi et al., 2016). This disease was also reported to be ‘common’ among individuals in Tierra del Fuego (Skewes pers. comm. in: Baldi et al., 2016). Hybridism with Llamas (L. glama), a species domesticated from guanacos has also been listed as a minor threat (Wilson and Mittermeier, 2011).
Whilst Baldi et al. (2016) remarked that harvesting of L. guanicoe for its meat for export had contributed to the value of the species and thereby “begun to reduce traditional conflict with sheep ranchers and foresters”,
15 Lama guanicoe
Moraga et al. (2014) reported that with the recovery of L. guanicoe numbers in southern Chile, the population in Tierra del Fuego was now in conflict with sheep ranching and commercial logging of the sub-Antarctic forest. In Tierra del Fuego, the majority of the population of L. guanicoe is distributed in the transition zone between forest and grassland (CONAMA, 2009). It was observed by Moraga et al. (2014) that the species demonstrated a preference for grassland, however, high sheep densities and range degradation exacerbated L. guanicoe’s use of forest habitats, with potential detrimental impacts on forest regeneration (as browsing may limit tree growth and reduce seedling density; Collado et al., 2009). Moraga et al. (2014) reported that a logging company in southern Tierra del Fuego had ‘leased the harvest of the species for sustainable use’ in the last three years, though the authors cautioned that such use may not necessarily reduce forest grazing if the spatial ecology of L. guanicoe and sheep management were not considered (Moraga et al., 2014).
Chile has 256 550 km² of protected areas3, (as of 2016), which do not cover the entire species’ range; the overlap was estimated at ~10% or 19 402 km² (Castillo et al., 2018). The species was reported to be present in 14 protected areas (national parks and reserves) within Chile’s National System of Protected Wild Areas (SNASPE): Lauca National Park (NP), Volcán Isluga NP, Llullaillaco NP, Pan de Azúcar NP, Nevado Tres Cruces NP, Llanos de Challe NP, Río Los Cipreses Nature Reserve (NR), Alto Bío-Bío NR, Cerro Castillo NR, Lago Jeinimeni NR, Lago Cochrane NP, Torres del Paine NP, Pali Aike NP, and Fray Jorge NP (CONAMA, 2009; Ministerio del Medio Ambiente, 2012; CONAF, 2013). Soto Volkart (2019) indicated that hunting under permit was not allowed within at least three National Parks or within a 4 km buffer zone outside of them.
L. guanicoe is protected by the National Hunting Law (Ley de Caza), which prohibits hunting of the species in any part of Chile with the following exemptions: for scientific purposes; to control populations causing significant damage to the ecosystem; to establish centres for captive breeding; and to allow for sustainable use of the resource, all of which require a permit from the relevant authority (Government of Chile, 2018). L. guanicoe has been legally hunted in Chile since 2003 (Baldi et al., 2016). Whilst Baldi et al. (2009) reported that the only hunted population was that of the island of Tierra del Fuego, maps outlining the zonation of guanaco hunting areas outside of Tierra del Fuego (for example, in the comunas of San Gregorio and Torres del Paine in Magallanes) were included in a presentation by SAG in 2019 (Soto Volkart, 2019), implying an increase in hunting areas.
In the early 1970s, the Chilean National Forest and Park Service (CONAF) established the Guanaco Conservation and Management Program in Magallanes region (Region XII) centred in Tierra del Fuego, and in 1980, CONAF created a national program to monitor and conserve key populations throughout Chile (Franklin, 1997). According to Baldi et al. (2016), there is no national management plan for L. guanicoe in Chile. However, a regional management plan was developed for northern and central Chile for 2010-2015 (Grimberg Pardo, 2010), and a plan was also reported to be in place for the Tierra del Fuego agricultural use area (see Tierra del Fuego section below; SAG Ministerio de Agricultura, 2016). The management plan for northern and central Chile aimed to address the lack of up-to-date distribution and abundance data for the species by carrying out further studies, as well as to improve monitoring and surveillance (to better tackle illegal hunting), and to increase or create protected areas that are important for the species (Grimberg Pardo, 2010). The plan also set out to implement management of feral dogs, one of the main threats to the species within the country (CONAF, 2013), and to encourage participation of rural and urban communities in conservation and education efforts in areas with L. guanicoe populations (Grimberg Pardo, 2010).
Tierra del Fuego: In the mid-1970s, the population of L. guanicoe in Tierra del Fuego had collapsed as a result of hunting pressure, habitat degradation, and competition with livestock (WCS, 2018). However, the population has since recovered; the number of individuals within the ‘Guanaco Conservation and Management Program’ area (also referred to as the ‘Guanaco Program area’ or program area) (central-south Tierra del Fuego in Chile) was reported to have increased from 5372 individuals in 1977 to 39 192 individuals in 2004
3 This estimate by Castillo et al. (2018) considers “all categories offering some level of protected” including private protected areas.
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(Table 7) (Morales, 2004 in: CONAMA, 2009). Téllez Millacari (2008) included the same population figures in a table labelled as “population of Guanacos in Tierra del Fuego 1977-2008”. The figures are the same up until 2003, where they diverge from the figures given by Morales (2004 in: CONAMA, 2009; Table 8). By 2010, the population of the Chilean part of Tierra del Fuego was estimated to have increased to between 183 000 and 197 000 individuals (Soto, 2010 in: González and Acebes, 2016). According to Soto (2004, in: Téllez Millacari 2008), densities of L. guanicoe in the program area in 2004 were close to the carrying capacity of the area; a high incidence of scabies lead the author to suggest that such high densities might also be compromising the health of the population. In a southern area of Tierra del Fuego (on Cameron ranch), population surveys carried out between 1977 and 2012 recorded 5372 individuals in 1977 and 40 548 in 2012 (Zubillaga et al., 2014b). While the first 25 years of data showed ‘apparent monotonic growth’, the population fluctuated over the last 10 years indicating that it had reached its carrying capacity (of ~45 000 individuals; Zubillaga et al., 2014b). In addition to population density (which varied between 2.7 and 30.7 ind./km2), the authors observed that average winter temperature and sheep densities were also significant in determining the population growth rate of L. guanicoe (Zubillaga et al., 2014b).
Table 7: Population of L. guanicoe in the Guanaco Program area (central-south of Tierra del Fuego). Source: Morales (2004) in: CONAMA (2009). Year Males Females Juveniles ‘Chulengos’ Unknown Total 1977 2487 677 0 381 1826 5372 1978 2395 1442 0 729 1178 5744 1979 4178 1312 0 250 1201 6940 1980 2510 890 0 207 3085 6693 1981 4746 1900 0 0 1651 8297 1982 4773 2232 1776 641 2911 12 334 1983 5399 2017 11 630 2614 10 670 1984 8471 3797 1412 1545 3854 19 078 1985 3635 2524 875 1088 3096 11 219 1987 5032 3272 644 1797 1933 12 393 1988 5393 3061 2006 1550 638 13 027 1989 5750 3284 1677 1818 1536 14 094 1990 7857 2629 350 1811 1183 14 604 1991 10 310 4444 533 1955 1778 17 775 1992 9265 4155 2555 3054 706 20 774 1993 7056 3282 1969 1969 2133 16 410 1994 9206 3160 941 1458 2861 17 626 1995 11 068 4598 1861 2076 1842 21 445 1997 11 190 4957 1394 3208 27 20 777 1998 15 616 6905 1698 3898 861 28 978 1999 14 021 7699 1012 4911 166 27 809 2000 10 660 8859 3360 5783 274 28 935 2001 15 919 8161 5055 5432 4275 38 842 2002 16 033 7886 4890 5472 4082 38 363 2003 13 471 7303 8638 4355 5635 39 403 2004 19 901 9434 3830 5689 338 39 192
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Table 8: Population of L. guanicoe in Tierra del Fuego, 2003-2008. Source: Téllez Millacari, 2008. Year Number 2003 32273 2004 43128 2005 58597 2006 52456 2007 33125 2008 61334
The recovery of L. guanicoe in Tierra del Fuego has been attributed to hunting restrictions implemented by the Government of Chile and a reduction in the number of sheep on the island (Baldi et al. 2009; WCS, 2018). It also led to the introduction of controlled culling of adults as a mechanism of population control and sustainable use (WCS, 2018), which has been in place since 2003 (Baldi et al. 2009).
Exempt Resolution No. 3.876 of 1 July 2013 delegated the power to authorize the hunting or capture of specimens of wild fauna, to allow the sustainable use of the resource in the cases indicated in the National Hunting Law, from the National Directorate of the SAG to Regional Directors (SAG Ministerio de Agricultura, 2014). Principles for the management of guanacos in the region are therefore set out in a region-specific management plan that is in turn approved by government resolutions (SAG Ministerio de Agricultura, 2016); however, this plan could not be accessed. Quotas appear to be approved on a company-by-company basis and are shared out between specific comunas, with hunting only allowed to take place in particular areas (Soto Volkart, 2019). Exempt Resolution No. 802/2014, for example, granted the Swanhouse company a quota or maximum hunting limit of 500 guanacos per year in the comuna of Primavera, and another 500 in the comuna of Porvenir (SAG Ministerio de Agricultura, 2014). However, because neither the management plan nor any legislation relating to the setting of quotas for particular years could be accessed, the mechanism currently used for setting quotas and the total area where hunting currently occurs are unclear. Hunting quotas have recently been issued for areas (e.g. Porvenir and Primavera; SAG Ministerio de Agricultura, 2014) that are outside of the ‘Guanaco program area’ for which the original SRG positive opinion was formed in 2003, but these are still within Tierra del Fuego.
The number of individuals authorised to be hunted in Tierra del Fuego has increased from 1700 in 2003 to 3750 in 2018 (Table 9). Between 2003 and 2018, quotas have been issued for 43 100 individuals in the Chilean Tierra del Fuego; 33 748 animals in total were reported to have actually been harvested (Soto Volkart, 2019).
Table 9: Quotas and actual number of L. guanicoe hunted in Tierra del Fuego, 2003-2018 (Soto Volkart, 2019). Numbers marked with an * have been estimated from the graph in Soto Volkart (2019) using a plot digitizer app4. Year Quota* Number Hunted 2003 1700 1700 2005 2000 2000 2006 2000 2000 2009 300* 77 2010 2000 1563 2011 3500 3500 2012 3500 3500 2013 4130* 3220 2014 4130* 3538 2015 2850* 1975
4 The web application https://automeris.io/WebPlotDigitizer/ allows the user to read values on a bar graph by specifying reference points on the graph’s axes, then measuring the height of each bar.
18 Lama guanicoe
Year Quota* Number Hunted 2016 5250* 1814 2017 6000 5476 2018 5760* 3750
In 2005, the quota for L. guanicoe was for 2000 adults two or more years old (of either sex), excluding females with young and juveniles less than a year old (SAG, 2005 in: Téllez Millacari, 2008). Téllez Millacari (2008) noted that the SAG had the ability to modify annual quotas in light of census data or unforeseen circumstances. Once extraction is completed, the Ministry of Agriculture (through CONAF and SAG) must conduct population monitoring in collaboration with the company or institution to which a quota is allocated (Téllez Millacari, 2008).
Hunting should preferentially take place between March and September, in order to minimise the impact of hunting on species reproduction (Téllez Millacari, 2008). Each application to hunt L. guanicoe must comply with the technical conditions of tender, which include, inter alia, providing information on the following (Téllez Millacari, 2008):
(1) The population structure (both historic and present) of the species in the area, and reproduction rates (2) The number, sex and age range of individuals to be hunted (3) Locations to be hunted (4) A detailed timetable of hunting activity planned (5) Arguments that demonstrate the project’s economic viability.
Culling in 2003 was reported to take place at five sites within which approximately 50% of the region’s guanaco population were found, killing 20% of local guanacos (Morales, 2004; Soto, 2005 in: Baldi et al., 2009). Previous quotas appear to have allowed hunters to take both sexes (e.g. SAG, 2005 in: Téllez Millacari, 2008).
Numerous challenges relating to the current system of harvest have been identified by the SAG, including the inclusivity of the current business model, difficulties in reaching the quota, and the perception by some stakeholders that current quotas are too low (Soto Volkart, 2019). In the short term, the SAG has planned to hold a workshop to discuss, inter alia (Soto Volkart, 2019):
i. Acceptable guanaco densities ii. Acceptable harvest models iii. How to standardise a simple, low-cost method for population monitoring. iv. How to revise current regulations to speed up the process of applying for a quota v. How to resolve the conflict between the quota potential, those interested in harvesting guanacos, and the capacity of harvest.
Parque Karukinka holds an L. guanicoe population that was identified as a priority population in Tierra del Fuego by the Regional Strategy for the Conservation of Guanacos (WCS, 2018). The park’s management plan aims to maintain densities of over 17 individuals/ha within both the park and within a 16 km area adjacent to the north of the park (WCS, 2018). The current harvesting model within the park was argued not to take account of variation in the ecology of guanacos, grasslands, and forest regeneration (WCS, 2018).
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en el norte de la Reserva La Payunia y su área de influencia (Mendoza, Argentina). Multequina (Latin American Journal of Natural Resources), 12: 37–48. Raedeke, K. 1979. Population dynamics and socioecology of the guanaco (Lama guanicoe) of Magallanes, Chile. University of Washington. 409 pp. Redford, K.H. and Eisenberg, J.F. 1992. Mammals of the Neotropics, the Southern cone: Chile, Argentina, Uruguay, Paraguay. University of Chicago Press, Chicago, USA. Rey, A., Novaro, A.J. and Guichón, M.L. 2012. Guanaco (Lama guanicoe) mortality by entanglement in wire fences. Journal for Nature Conservation, 20(5): 280–283. Rivas, L.F., Novaro, A.J., Funes, M.C. and Walker, R.S. 2015. Rapid assessment of distribution of wildlife and human activities for prioritizing conservation actions in a Patagonian landscape. PLoS ONE, 10(6). SAG 2005. Anexo de contrato de Licitacion cuota de extraccion de guanacos en isla Tierra del Fuego para la implementacion de criaderos, el desarollo de proyectos de repoblamiento y/o proyectos de aprovechamiento sustentable. SAG Ministerio de Agricultura 2016. Resolucion Extenta No. 273/2016. SAG Ministerio de Agricultura 2014. Resolucion Extenta No. 802/2014. Sarno, R.J., Clark, W.R., Bank, M.S., Prexl, W.S., Behl, M.J., Johnson, W.E. and Franklin, W.L. 1999. Juvenile guanaco survival: Management and conservation implications. Journal of Applied Ecology, 36(6): 937– 945. Schroeder, N.M., Matteucci, S.D., Moreno, P.G., Gregorio, P., Ovejero, R., Taraborelli, P. and Carmanchahi, P.D. 2014. Spatial and seasonal dynamic of abundance and distribution of guanaco and livestock: Insights from using density surface and null models Festa-Bianchet, M. (Ed.). PLoS ONE, 9(1): e85960. Schroeder, N.M., Ovejero, R., Moreno, P.G., Gregorio, P., Taraborelli, P., Matteucci, S.D., Carmanchahi, P.D., Leal, R. and Waterman, J. 2013. Including species interactions in resource selection of guanacos and livestock in Northern Patagonia. Sielfeld, W. 2004. Consultoría para el estudio poblacional de guanacos y tarucas asociados a la producción agropecuaria de la precordillera de la provincia de Parinacota. Universidad Arturo Prat y SAG. Sosa, R.A. and Sarasola, J.H. 2005. Habitat use and social structure of an isolated population of guanacos (Lama guanicoe) in the Monte Desert, Argentina. European Journal of Wildlife Research, 51(3): 207–209. Soto, F.V. 2014. Situación de Lama guanicoe en el Parque Nacional Llanos de Challe y su potencial como producto turístico. Revista Interamericana de Ambiente y Turismo - RIAT, 10(2): 181–188. Soto, N. 2004. Minuta uso sustentable del guanaco en Isla Tierra del Fuego. Santiago, Chile. 8 pp. Soto, V.N. 2010. Distribución y abundancia de la población de guanacos (Lama guanicoe, Muller 1776) en el área agropecuaria de Tierra del Fuego (Chile) y su relación de carga animal con la ganadería ovina. Universidad Internacional de Andalucia, Punta Arenas, Chile. Soto, V.N. 2005. Soto Volcart, N. (2005) Evaluacion del primer año de caza, monitoreo de la poblacion 2004 y determinacion de cuota de extraccion año 2005. Report to Servicio Agricola y Ganadero, Chile. Soto Volkart, N. 2019. Guanaco en Magallanes: contexto y marco normativo. Téllez Millacari, L. 2008. Proceso de extraccion y aprovechamiento sustentable del guanaco (Lama guanicoe, Müller) mediante cuotas de caza en Tierra del Fuego. 47 pp. von Thüngen, J. and Lanari, M.R. 2010. Profitability of sheep farming and wildlife management in Patagonia. Pastoralism, 1(2): 274–290. Travaini, A., Zapata, S.C., Bustamante, J., Pedrana, J., Zanón, J.I. and Rodríguez, A. 2015. Guanaco abundance and monitoring in southern patagonia: Distance sampling reveals substantially greater numbers than previously reported. Zoological Studies, 54(JAN): 0–12. Valladares Faúndez, P. 2012. Mamíferos terrestres de la Región de Atacama, Chile: Comentarios sobre su distribución y estado de conservación. Gayana (Concepción), 76(1): 22–37. WCS 2018. Plan de Manejo: Parque Karukinka. Tierra del Fuego, Chile. 146 pp. Wilson, D.E. and Mittermeier, R.A. (Eds.) 2011. Handbook of the mammals of the world: Vol. 2. Hoofed mammals. Lynx editions, Barcelona. Wilson, D.E. and Reeder, D.M. 1993. Mammal species of the world: a taxonomic and geographic reference. 2nd Ed. Smithsonian Institution Press, Washington DC. 1207 pp. Yahnke, C.J., Gamarra De Fox, I. and Colman, F. 1998. Mammalian species richness in Paraguay: The effectiveness of national parks in preserving biodiversity. Biological Conservation, 84(3): 263–268. Zubillaga, M., Skewes, O., Soto, N. and Rabinovich, J.E. 2014a. Density but not climate affects the population growth rate of guanacos (Lama guanicoe) (Artiodactyla, Camelidae). F1000Research, 2: 210. Zubillaga, M., Skewes, O., Soto, N., Rabinovich, J.E. and Colchero, F. 2014b. Bayesian inference on the effect of density dependence and weather on a Guanaco population from Chile. PloS one, 9(12): e115307.
22 Crocodylus niloticus
CROCODYLIA: CROCODYLIDAE
Crocodylus niloticus II/B
SYNONYMS: Alligator cowieii Smith, 1937; Crocodilus binuensis Baikie, 1857; Crocodilus chamses Bory, 1824; Crocodilus complanatus Geoffroy, 1827); Crocodilus lacunosus Geoffroy, 1827; Crocodilus madagascariensis Grandidier, 1872; Crocodilus marginatus Geoffroy, 1827; Crocodilus multiscutatus Rüppell, 1826; Crocodilus octophractus Rüppell, 1831; Crocodilus robustus Vaillant & Grandidier, 1872; Crocodilus suchus Geoffroy, 1807; Crocodilus vulgaris Cuvier, 1807
COMMON NAMES: Nile Crocodile (EN), Crocodile du Nil (FR), Cocodrilo del Nilo (ES)
RANGE STATES: Angola, Benin, Botswana, Burkina Faso, Burundi, Cameroon, Central African Republic, Chad, Congo, Côte d'Ivoire, Democratic Republic of Congo, Egypt, Equatorial Guinea, Eritrea, Eswatini, Ethiopia, Gabon, Gambia, Ghana, Guinea, Guinea Bissau, Kenya, Liberia, Madagascar, Malawi, Mali, Mauritania, Mozambique, Namibia, Niger, Nigeria, Rwanda, Senegal, Sierra Leone, Somalia, South Africa, South Sudan, Sudan, Togo, Uganda, United Republic of Tanzania, Zambia, Zimbabwe
UNDER REVIEW: Malawi
EU DECISIONS: Current positive opinion for wild-taken specimens from Namibia and ranched specimens from Mozambique formed on 05/02/2018
Current positive opinion for hunting trophies and ranched specimens from Zimbabwe formed on 27/06/2016. Current no opinion i) for wild-sourced specimens from Zimbabwe, with the exception of hunting trophies, formed on 27/06/2016
Current no opinion ii) for wild-sourced specimens from Madagascar first formed on 30/01/2015 and last confirmed on 09/04/2015
IUCN: Least Concern
Taxonomic note
Whilst previously considered a subspecies of Crocodylus niloticus, Schmitz et al. (2003) proposed recognising Crocodylus suchus as a distinct species based on genetic analysis. C. suchus is currently recognised as a valid species by the IUCN Red List (Isberg et al., 2019) and in the online reptile database (Uetz et al., 2019). In a CoP18 document relating to standard nomenclature for animals (CoP18 Doc. 99 Annex 6), it is proposed that C. suchus is recognised as an accepted species within CITES in accordance with Hekkala et al. (2011). C. niloticus [sensu stricto] was noted to occur in eastern Africa, the Nile valley and southern Africa (and therefore includes Malawi), with C. suchus occurring in west and central Africa, including the Congo, Niger and Ogoue river drainages (Shirley et al., 2015; Isberg et al., 2019). Trade patterns
Crocodylus niloticus was listed in Appendix I on 01/07/1975. Several countries (including Malawi) were included in Appendix II on 01/08/1985 subject to quotas. On 18/01/1990, the population of Malawi was
23 Crocodylus niloticus maintained in Appendix II without being subject to specified annual export quotas. C. niloticus was listed in Annex A of the EU Wildlife Regulations on 01/06/1997, except for populations included in Appendix II including those of Malawi, which were included in Annex B.
CITES annual reports have been submitted by Malawi for all years 2008-2017. Malawi published export quotas for ranched and wild-sourced C. niloticus skins 2008-2012 (Table 1).
Table 1: CITES export quotas for ranched and wild-sourced Crocodylus niloticus from Malawi, 2008-2017.
2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 2018 Quota (ranched skins) 3000 3000 3000 10000 5000 Quota (wild-sourced skins) 200 200 200 200 150
According to the CITES Trade Database, direct trade in C. niloticus from Malawi to the EU-28 2008-2017 predominantly comprised ranched skins for commercial purposes, totalling 18 760 as reported by Malawi, and 13 084 as reported by importers. Trade in ranched skins from Malawi to the EU-28 increased by more than six times in 2017 compared to 2016 trade levels, as reported by EU importers. Malawi reported a six-fold increase in trade to the EU in 2016 compared to 2015, while trade in 2017 remained at a comparable level (Table 2). Lower quantities of wild-sourced skins for commercial purposes were reported, but decreased over the ten years (Table 2).
Direct trade from Malawi to the rest of the world 2008-2017 predominantly comprised ranched skins for commercial purposes, with relatively consistent trade levels, although exports generally increased from 2011 to a peak in 2016.
Indirect trade originating in Malawi to the EU-28 comprised largely of ranched small leather products and skins for commercial purposes, with lower quantities of wild-sourced small leather products for commercial purposes (Table 3).
24 Crocodylus niloticus
Table 2: Direct exports of Crocodylus niloticus from Malawi to the EU-28 (EU) and the rest of the world (RoW), 2008-2017. Due to low levels of trade (<10 units), trade in meat, trophies and skulls have been excluded from the table. Importer Term Purpose Source Reported by 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 Total EU leather products (small) T R Importer 14 2 16 Exporter skin pieces H W Importer 1 1 Exporter T R Importer Exporter 443 443 skins H W Importer 1 4 2 7 Exporter 1 1 P W Importer Exporter 399 399 T R Importer 1667 600 400 1002 537 1064 1243 882 5689 13084 Exporter 3170 2403 200 1114 922 5916 5035 18760 W Importer 200 200 100 96 100 100 2 2 800 Exporter 200 200 400 RoW leather products (large) P I Importer 1 1 Exporter T R Importer 504 504 Exporter leather products (small) T C Importer 1000 1000 Exporter R Importer 5320 5320 Exporter skin pieces T R Importer 150 150 Exporter 1000 1000 skins H W Importer Exporter 1 1 P I Importer Exporter 1 1 T C Importer 500 250 750 Exporter R Importer 1158 2402 3173 5505 4092 4345 20675 Exporter 1308 4949 5371 2784 5324 6065 25801 W Importer 10 10 Exporter 2 116 414 532
25 Crocodylus niloticus
Importer Term Purpose Source Reported by 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 Total trophies H W Importer 2 1 3 Exporter Source: CITES Trade Database, UNEP-WCMC, Cambridge, UK, downloaded on 04/02/2019.
Table 3: Indirect exports of Crocodylus niloticus originating in Malawi to the EU-28, 2008-2017. Term Purpose Source Reported by 2008 2009 2010 2011 2012 2013 2014 2015 2016 2017 Total leather products (large) Q R Importer Exporter 7 7 T R Importer Exporter 6 6 leather products (small) Q R Importer Exporter 15 15 T C Importer 14 14 Exporter R Importer 5 93 5 27 58 48 68 304 Exporter 2 9 106 1 21 40 59 76 27 341 W Importer 10 3 3 13 39 2 70 Exporter 10 4 5 12 39 70 - - Importer 3 3 Exporter skins T R Importer 20 25 62 107 Exporter 20 25 112 157 Source: CITES Trade Database, UNEP-WCMC, Cambridge, UK, downloaded on 04/02/2019.
26 Crocodylus niloticus
Conservation status
Crocodylus niloticus was reported to be the most widely distributed crocodile species in Africa (Fergusson, 2010) and is found within 26 countries in eastern, cental and southern Africa (Fig. 11; Isberg et al., 2019). It occurs throughout all sub- Saharan Africa, except for southwestern Africa, and extends up through northeastern Africa to Egypt (Fergusson, 2010; Isberg et al., 2019). It is also present on Madagascar (Fergusson, 2010; Isberg et al., 2019). Historical records show that C. niloticus [sensu lato] once ranged over a much greater extent, but is now extirpated from Algeria, Comoros Islands, Israel (Trutnau and Sommerlad, 2006; Fergusson, 2010) and Djibouti (Groombridge, 1982). In East Africa the species has been recorded from sea level to an altitude of 1600 m (Spawls et al., 2018).
C. niloticus occurs in a wide range of wetlands habitats including rivers, lakes and swamps and coastal estuaries (Thorbjarnarson et al., 1992; Trutnau and Sommerlad, 2006; Isberg et al., 2019). As an apex predator, C. niloticus has been reported to be an indicator of the health of the Figure 1: Range of Crocodylus niloticus. Source: Isberg ecosystems in which it occurs (Ashton, 2010). et al. (2018).
C. niloticus is among the largest and most biologically studied crocodile species (Fergusson, 2010). Adults typically reach lengths of 2-3.5 m (Spawls et al., 2018), however, this species is sexually dimorphic and in exceptional cases males have been recorded to reach lengths of up to 6 m (Thorbjarnarson et al., 1992; Fergusson, 2010). Adults of C. niloticus [sensu lato] reach sexual maturity around 12-15 years of age (Trutnau and Sommerlad, 2006), with females reported to sexually mature between 1.8 m (Games, unpublished data in: Fergusson, 2010) and 3 m in length (Detouf-Boulade, 2006 in Fergusson, 2010). Clutch size was reported to vary considerably among C. niloticus populations (Fergusson, 2010). According to Spawls et al. (2018) females typically lay between 30-50 eggs, with extremes of 20 and 95 eggs noted. Females were reported to incubate the eggs for a period of 75-90 days and to guard the young for 6-8 weeks after hatching (Fergusson, 2010; Combrink et al., 2016). C. niloticus eggs were reported to be predated by monitor lizards (Combrink et al., 2011, Calverley and Downs, 2017), marsh mongoose (Combrink et al., 2016), hyaenas and humans (Isberg et al., 2019).
C. niloticus populations were reported to have been historically depleted due to hunting across much of the range between the 1940s and 1960s (and as late as 1970s in some countries; Isberg et al., 2019). However, as a result of increased levels of protection through national legislation and international trade regulations (e.g. CITES), populations were noted to have recovered throughout many parts of the range (Fergusson, 2010).
C. niloticus was categorised globally as Least Concern by the IUCN Red List on the basis that, despite some evidence of localised population declines, the species was considered to generally remain widespread and stable, and was not threatened with extinction (Isberg et al., 2019). The Red List assessment provided no
1 Disclaimer: The boundaries and names shown and the designations used on maps do not imply official endorsement or acceptance by UN Environment or contributory organisations.
27 Crocodylus niloticus overall global population size estimates. In an non-detriment finding report on the species specifically for Kenya, the global population of C. niloticus in the wild was estimated at 250 000 to 500 000 individuals (Kyalo, 2008), however no information was provided on how this estimate was reached. Populations of C. niloticus in West Africa (now considered by the IUCN to be C. suchus), were considered to be depleted and less numerous than C. niloticus [sensu stricto]. An IUCN assessment for C. suchus is currently in preparation and not yet available, but it was considered by IUCN experts that despite C. suchus appearing to meet the historical decline criteria for the Endangered category (over the past three generations), the species was not Endangered, and the actual probability of extinction was low (IUCN Crocodile Specialist Group, 2019).
The population status of C. niloticus across range States is variable. Recent surveys suggested a stable adult population of C. niloticus in Botswana’s Okavango panhandle, but with a decline in sub-adults, possibly as a result of reduced fish availability (Bourquin and Shacks, 2016), a “healthy” population in Kenya in areas under control of the Kenyan Wildlife Service (Haller pers. comm. 2016 in: Isberg et al., 2019) and a stable population in the Murchison Falls National Park in Uganda (Behangana et al., 2017). Four large “possibly secure” populations were noted to occur in South Africa (Isberg et al., 2019). A population increase was reported in Egypt, where the species was noted to be recovering (Shirley et al., 2012), and in Namibia, increases were reported from the Okavango River (Du Preez et al., 2014). The species was reported to be still found in large numbers in major rivers in Zambia (Thomas and Reilly pers. comm. in: Isberg et al., 2019). While no recent population estimates were available for Tanzania, Mwita and Games (2012) noted that the Selous Game Reserve had “one of the most impressive Nile crocodile populations in Africa”. Surveys of the Zambezi River (which flows through Zambia, Angola, Namibia, Botswana, Zimbabwe and Mozambique) in 2006 and 2009 indicated that C. niloticus populations had increased over this period (Wallace et al., 2013).
However, declines have been reported in the central African countries of Cameroon, Gabon and Central African Republic (Isberg et al., 2016), as well as Ethiopia (Shirley et al., 2014) and Madagascar (Ottley et al., 2008). Populations in Zimbabwe were reported to have increased in park estates, but declines were reported in communal and forestry Lands (Crocodile Farmers Association of Zimbabwe, 2015 in: Isberg et al., 2019). No population data was reported to be available for other range States (Isberg et al., 2019).
C. niloticus has a history of over-exploitation stemming from the popularity of the species in the leather industry and its prominence as a trophy species (Hekkala et al., 2010). Raw skins remain the most significant product in global trade, with 201 000 C. niloticus skins reported exported 2006-2015 on average (Caldwell, 2017). Fergusson (2010) considered that the illegal trade in C. niloticus was insignificant in southern and East Africa, and the establishment of the formal crocodile skin industry and ongoing trade controls had largely negated illegal trade in skins. Hekkala et al. (2010) noted that while C. niloticus populations in East Africa had “apparently recovered,” the species’ range remained fragmented due to a range of threats including increasing aridity and anthropogenic land-use change. Zisadza-Gandiwa et al. (2013) reported declines in certain C. niloticus populations were likely to continue due to habitat destruction and anthropogenic factors (human- wildlife conflict, poaching, commercial utilisation and pollution). Moshoeu (2017) also noted that C. niloticus was the most sought after reptile species for use in therapeutic purposes in Africa, with trade recorded in 17 countries including Malawi.
Threats to the species were reported to be increasing outside of formally protected areas, and in some cases, within them (Combrink et al., 2011; Calverley and Downs, 2017; Behangana et al., 2017). Aust (2009) noted that crocodile densities were significantly negatively correlated with human densities and positively correlated with protected areas; similarly Wallace et al. (2013) noted lower densities in areas of increased human presence and greater densities in areas of increased wildlife and habitat protection. As a result of escalating human-crocodile conflict, retaliation killings were noted to have become a threat to the species in certain areas (Ross, 1998; Isberg et al., 2019).
Ross (1998) noted that national laws protecting C. niloticus and international trading regulations had resulted in the recovery of the species in many parts of its range. Whilst many management initiatives have been aimed
28 Crocodylus niloticus at implementing sustainable management programs in countries that utilized crocodiles (Fergusson, 2010), Isberg et al. (2019) noted that there had been a shift towards the management of human-crocodile conflicts within several countries, replacing the emphasis on sustainable use programmes and trophy hunting. Aust (2009) noted that the future status of national populations would be linked to the extent and management of nationally protected habitat.
Malawi: A study conducted in the mid-1980s indicated that C. niloticus was widely distributed across Malawi (Thorbjarnarson et al., 1992). Based on Isberg et al. (2019), the species’ occurrence within the country is depicted in Fig. 1 (and inset). C. niloticus has been recorded from Elephant Marsh on the Lower Shire River in Malawi (Kosamu et al., 2012), Lake Chilwa (Nagoli et al., 2016), Liwonde National Park bordering the Shire River and Lake Malombe (Hutton 1989 in Thorbjarnarson et al., 1992; Leroux and Reid, 2016 in: Isberg et al., 2019), Salima Bay on Lake Malawi (Hekkala et al., 2010) and Boadzulu Island on Lake Malawi (Government of Malawi, 2018).
No national population estimates for Malawi could be located. However, large C. niloticus populations were noted to occur in the Liwonde National Park and Elephant Marsh (Hutton 1989 in: Thorbjarnarson et al., 1992). In Liwonde National Park, 676 crocodiles, measuring above one metre in length were counted by aerial surveys along a 38 km stretch of the Shire River, from the boundary of Lake Malombe in 2016, equating to a density of 17.8 crocodiles/km (Leroux and Reid, 2016 in Isberg et al., 2019); this was reported as one of highest densities recorded in Africa (IUCN SCG Action Plan for C. niloticus, in prep.).
A previous survey carried out across a 184 km stretch of the lower reaches of the Lower Shire River between Kapichira Falls and Nsanje in 2005 observed 317 crocodiles (density of 1.72 crocodiles/km), from which an estimate of 745 crocodiles were estimated in this stretch (Fergusson, 2005 in: IUCN SCG Action Plan for C. niloticus, in prep.). Based on an estimate of 1222 crocodiles within an un-surveyed section of the same river (320 km), a total figure of 1967 crocodiles within the entire 504 km river section was estimated at that time (Fergusson, 2005 in: IUCN SCG Action Plan for C. niloticus, in prep.). It was reported that adult densities were significantly lower in 2005 than in previous years, with adult densities declining from north to south, suggesting higher offtake in the latter (Fergusson, 2005 in: IUCN SCG Action Plan for C. niloticus, in prep.). Records of Fergusson’s individual surveys in 2005 on the Lower Shire river in Malawi, as well as records from 1997 were found within the Crocodile Survey Database (Hutton and Lainez, 2010); these show a decline in densities of C. niloticus from 1997 to 2005 within specific stretches (Table 4). All historical records in Table 4 were much lower than the densities found in 2016 in the Liwonde National Park by Leroux and Reid (2016).
No information was reported to be available on C. niloticus numbers in Lake Malawi, although crocodile attacks had been documented along the south, north and western shoreline (IUCN SCG Action Plan for C. niloticus, in prep.).
Table 4: C. niloticus survey data for Lower Shire River, Malawi based on records in the Crocodile Survey Database (Hutton and Lainez, 2010). [Densities presumed to be individuals/km, but no units provided]. Location Total count Density Year Source Chikwawa bridge to 138 2.99 1997 Bhima and Jamusana (1998) Hippo and crocodile counts in Sucoma the Lower Shire river. Report submitted to IUCN Zambezi Basin Shiromo to Sucoma 313 5.28 1997 As above Chiromo bridge to Nsanje 195 3.12 1997 As above Kapichira to Chikwawa 45 2.78 2005 Fergusson (2005) Survey of crocodile populations and bridge crocodile/human conflict on the lower Shire river, Malawi
Chikwawa bridge to 104 2.25 2005 As above Sucoma Sucoma to Chiromo 91 1.53 2005 As above bridge Chiromo bridge to Nsanje 77 1.23 2005 As above
29 Crocodylus niloticus
Human-crocodile conflict is a problem in Malawi, and according to the worldwide crocodilian attack database, 689 fatal and 292 non-fatal C. niloticus attacks on humans were reported 2008-2019 (CrocBITE, 2019). However, Isberg et al. (2019) noted that attack reports across the range States of C. niloticus represent only a subset of actual attacks, as some will go unreported. According to a Malawian medical journal, around 800 crocodiles were reported to be culled annually in the country as a result of human-crocodile conflict (Wamisho et al., 2009). The species was also reported to be hunted by both government hunters and private concessionaires; offtake in 2005 was considered to be 103 crocodiles on the Lower Shire River, which at that time, was not thought to be sustainable (Fergusson, 2005 in: IUCN SCG Action Plan for C. niloticus, in prep.).
Crocodile ranching and farming of captive-bred individuals in Malawi has been carried out on a small scale since the 1980s (AC22 Inf. 2). Three C. niloticus farms were noted to exist in the country in 2003 (Salima-Nyika Crocodile Farm, Koma Crocodile Farm and Crocodile Farming & Research Centre; AC22 Inf. 2). At that time, it was noted that the ranching programme in Malawi was moribund, with any remaining activities poorly recorded (AC22 Inf. 2). The Government of Malawi (2010) reported that four ranching facilities were noted to exist in the country in 2008-2009. The number of C. niloticus individuals and eggs within these facilities at that time are summarised in Table 5. According to the National Parks and Wildlife (Wildlife Ranching) Regulations of 1994, a person operating a wildlife ranch on which crocodiles were raised should make available to the Department of National Parks and Wildlife at least 10% of the hatchlings of each egg collection effort after rearing them to a length of at least one metre (Government of Malawi, 1994a).
Crocodile farms had reportedly been authorised to capture 30 live adults annually for five years, with a five year permit to collect 5000 eggs from the wild (Government of Malawi, 2010). It was also noted that in all farms except Koma, local community sensitisation programmes were planned for 2010, whereby farmers would receive payment for 5% of surviving crocodiles from wild eggs to be given to National Parks and Wildlife, rather than releasing them locally (Government of Malawi, 2010). According to the website of Shire River Crocodiles Ltd2, which has two farms located in Kaombe and Ngabu in Malawi, the company holds 18 000 individuals of C. niloticus, maintaining its own breeding stock, with harvests of crocodiles at an average age of three years. The website also reported that the company “enjoys Department of National Parks & Wildlife of Malawi Sanctuary Status”.
Table 5: Details of ranching facilities for C. niloticus in Malawi in 2008 and 2009 (Government of Malawi, 2010). Company name Location No. Breeding Wild eggs Eggs Farmed Exports of skins crocodiles stock collected collected eggs (year) hatched Nyika Crocodile Farm Salima 13 635 78 1700 2136 1500 (2008) (2008) (2008) 1503 (2009) 1300 0 (2009) (2009) Shire Valley Chikhwawa 4339 2544 (2008) 2089 First export was Crocodiles (Ngabu) 3232 (2009) (2008) expected in 2011 Farm 2561 (2009) Chiwale Crocodile Thyolo 4328 2212 (2008) First export was Farm 2374 (2009) expected in 2011 Koma Croc Farm Mangochi 655 69 Not progressing with export
Webb (pers. comm. to UNEP-WCMC, 2019) stated that the global crocodile industry had shown a recent downturn with supply outstripping demand, and this was likely to have reflected on ranching facilities in Malawi. The IUCN Crocodile Specialist Group (CSG) (in litt. to UNEP-WCMC, 2019) considered that overall, it was likely that the limited off-take of C. niloticus for ranching and subsequent export would not have a detrimental effect on the conservation of the species in Malawi.
2 http://www.shirerivercrocodiles.com/ [Accessed 7/6/2019]
30 Crocodylus niloticus
The conservation of selected wildlife in Malawi is stipulated by the National Parks and Wildlife Act (Act No. 11 of 1992) and its amendment (Government of Malawi, 1992; 2017). C. niloticus is listed as a protected species, as is any reptile within a national park or wildlife area (Government of Malawi, 1994b). Special licences can be issued in respect of protected species to authorise hunting for the purposes of scientific research, educational or “other proposer use” (e.g. zoos, museums), and hunting licences can be authorised to hunt wild protected species outside of protected areas, or inside protected areas subject to a special permit by the department’s director (provided certain conditions are met) (Government of Malawi, 2017).
Hunting of C. niloticus was reported to be permitted under tourist hunting concession trials outside of protected areas (Macpherson, 2013 in: Isberg et al., 2019). The Government of Malawi was reported to have received revenue of USD 9796 in 2012 from hunting 16 individuals of C. niloticus by eight hunters (Macpherson, 2013 in: Isberg et al., 2019). Two protected areas (Elephant Marsh and Lake Chilwa) that were reported to contain C. niloticus populations are Ramsar wetland sites of international importance, and Lake Malawi National Park is a World Heritage Site (UNEP-WCMC and IUCN, 2019); the species was additionally reported to occur in the Liwonde National Park (Hutton 1989 in Thorbjarnarson et al., 1992).
It was noted that a crocodile management plan for Malawi was published by FAO in 1990, but it had not been updated in 2004 (AC22 Inf. 2). References
Ashton, P.J. 2010. The demise of the Nile crocodile (Crocodylus niloticus) as a keystone species for aquatic ecosystem conservation in South Africa: The case of the Olifants River. Aquatic Conservation: Marine and Freshwater Ecosystems, 20(5): 489–493. Aust, P.W. 2009. The ecology, conservation and management of Nile crocodiles Crocodylus niloticus in a human dominated landscape. Imperial College London. 148 pp. Behangana, M., Lukwago, W., Dendi, D., Luiselli, L. and Ochanda, D. 2017. Population surveys of Nile crocodiles (Crocodylus niloticus) in the Murchison Falls National Park, Victoria Nile, Uganda. European Journal of Ecology, 3(2): 67–76. Bishop, J.M., Leslie, A.J., Bourquin, S.L. and O’Ryan, C. 2009. Reduced effective population size in an overexploited population of the Nile crocodile (Crocodylus niloticus). Biological Conservation, 142(10): 2335–2341. Bourquin, S.L. and Shacks, V.A. 2016. Population trends in a previously exploited Nile crocodile population in the Okavango panhandle, Botswana. Crocodile Specialist Group Newsletter, 35(4): 5–6. Calverley, P.M. and Downs, C.T. 2017. The Past and Present Nesting Ecology of Nile Crocodiles in Ndumo Game Reserve, South Africa: Reason for Concern? Journal of Herpetology, 51(1): 19–26. Caldwell, J. 2017 World trade in crocodilian skins 2013-2015. UNEP-WCMC, Cambridge. 32 pp. Combrink, X., Korrûbel, J.L., Kyle, R., Taylor, R. and Ross, P. 2011. Evidence of a declining Nile crocodile (Crocodylus niloticus) population at Lake Sibaya, South Africa. South African Journal of Wildlife Research, 41(2): 145–157. Combrink, X., Warner, J.K. and Downs, C.T. 2016. Nest predation and maternal care in the Nile crocodile (Crocodylus niloticus) at Lake St Lucia, South Africa. Behavioural Processes, 133(January 2018): 31–36. Crocodile Farmers Association of Zimbabwe 2015. The performance of crocodile farming in Zimbabwe. Report submitted at the 40th Annual Meeting between CFAZ and the Parks and Wildlife Management Authority. Crocodile Specialist Group (CSG) 2019. CSG pers. comms. to UNEP-WCMC, 3rd June 2019. Fergusson, R.A. 2010. Nile Crocodile Crocodylus niloticus. In: Manolis, S.C. and Stevenson, C. (Eds.). Crocodiles. Status Survey and Conservation Action Plan. Crocodile Specialist Group, Darwin. 84–89. Fergusson, R.A. 2005. Survey of crocodile popualtions and crocodile/human conflict on the lower Shire River, Malawi. Government of Malawi, 1992. Malawi National Parks and Wildlife Act (1992). Available at: https://www.ecolex.org/details/legislation/national-parks-and-wildlife-act-act-no-11-of-1992-lex- faoc004733/ [Accessed 10/6/2019] Government of Malawi, 1994a. National Parks and Wildlife (Wildlife Ranching) Regulations, 1994. Government Notice No. 82. Government of Malawi, 1994b. National Parks and Wildlife (Protected Species) (Declaration) (Order), 1994.
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Government Notice No. 89. Government of Malawi, 2010. Report on crocodile (Crocodylus niloticus) ranching in Malawi for 2008 and 2009. CITES National Report on ranching operations in Malawi. Available at: https://www.cites.org/sites/default/files/common/resources/reports/ranch/MW1005.pdf [Accessed 6/6/2019] Government of Malawi, 2017. Act to amend the National Parks and Wildlife Act. Available at: http://wildlife.gov.mw/wp-content/uploads/2018/04/NPW-Act-2017.pdf [Accessed 10/6/2018]. Government of Malawi, 2018. State of conservation report by the State Party. World Heritage Convention Report. Available at: https://whc.unesco.org/document/167399 [Accessed 10/6/2019]. Groombridge, B. 1982. The IUCN amphibia-reptilia Red Data Book Part 1: Testudines, Crocodylia, Rhynchocephalia. IUCN, Gland, Switzerland. Hekkala, E., Shirley, M.H., Amato, G., Austin, J.D., Charter, S., Thorbjarnson, J., Vliet, K.A., Houck, M.L., Desalle, R. and Blum, M.J. 2011. An ancient icon reveals new mysteries: mummy DNA resurrects a cryptic species within the Nile crocodile. Molecular Ecology, 20(20): 4199–4215. Hekkala, E.R., Amato, G., DeSalle, R. and Blum, M.J. 2010. Molecular assessment of population differentiation and individual assignment potential of Nile crocodile (Crocodylus niloticus) populations. Conservation Genetics, 11: 1435–1443. Hutton, J. and Lainez, D. 2010. The Crocodile Survey Database. http://www.crocsurveys.net/Home [Accessed 6/6/2019] IUCN CSG Action Plan for C. niloticus (in prep.). IUCN Crocodile Specialist Group in litt. to UNEP-WCMC, 3rd June 2019. Isberg, S., Combrink, X., Lippai, C. and Balaguera-Reina, S. 2019. Crocodylus niloticus. The IUCN Red List of Threatened Species 2019 https://www.iucnredlist.org [Accessed 20 May 2019]. Isberg, S.R., McLeod, R. and Ross, P. 2016. Crocodylus niloticus Red List workshop report. Report to the Steering Committee of the 24th Working Meeting of the IUCN-SSC Crocodile Specialist Group, Skukuza, South Africa 23-26th May, 2016. Kosamu, I., de Groot, W., Kambewa, P. and de Snoo, G. 2012. Institutions and Ecosystem-Based Development Potentials of the Elephant Marsh, Malawi. Sustainability, 4(12): 3326–3345. Kyalo, S. 2008. Non-detriment finding studies on Nile crocodile (Crocodylus niloticus) the status of and trade in the Nile Crocodile in Kenya. In: NDF Workshop Case Studies, WG7 - Reptiles and Amphibians. Available at: https://cites.unia.es/file.php/1/files/WG7-CS1.pdf. [Accessed 1/6/2019] Leroux, P. and Reid, C. 2016. Liwonde National Park aerial census dry season August. Unpublished internal report. Macpherson, D. 2013. Report on the fourth consecutive year of trial tourist hunting in Malawi. Kanongo Estate, Malawi. Unpublished report. 4 pp. Moshoeu, T.T.J. 2017. Overview of the trade of reptile taxa consumed for therapeutic purposes across Africa. University of Witwatersrand, South Africa. 102 pp. Mwita, M. and Games, I. 2012. Tanzanian crocodile survey. Report to the Director of Wildlife, Tanzania. Nagoli, J., Chiwona-Karltun, L., Likongwe, P., Mulwafu, W. and Green, E. 2016. Conflicts over Natural Resource Scarcity in the Aquatic Ecosystem of the Lake Chilwa. Environment and Ecology Research, 4(4): 207–216. Ottley, B., Lippai, C. and Rakotondrazafy, A.M.A. 2008. Surveys of wild crocodile populations in Madagascar. Final report to GTZ. Du Preez, P., Beytell, P.C., Louw, C. and Lyet, A. 2014. Aerial survey of the Nile crocodile population in the Okavango River from Angola boundary to Botswana boundary October 2013. Report to the Directorate Scientific Services, Ministry of Environment and Tourism, Namibia. Ross, J.P. 1998. Crocodiles: Status survey and conservation action plan. IUCN/SSC Crocodile Specialist Group. IUCN, Gland, Switzerland and Cambridge, UK. Schmitz, A., Mansfield, P., Hekkala, E., Shine, T., Nickel, H., Amato, G. and Böhme, W. 2003. Molecular evidence for species level divergence in African Nile Crocodiles Crocodylus niloticus (Laurenti, 1786). Comptes Rendus Palevol, 2(8): 703–712. Shirley, M.H., Dorazio, R.M., Abassery, E., Elhady, A.A., Mekki, M.S. and Asran, H.H. 2012. A sampling design and model for estimating abundance of Nile crocodiles while accounting for heterogeneity of detectability of multiple observers. The Journal of Wildlife Management, 76(5): 966–975. Shirley, M.H., Oduro, W. and Beibro, H.Y. 2009. Conservation Status of Crocodiles in Ghana and Cote-d’Ivoire, West Africa. Oryx, 43(1): 136–145. Shirley, M.H., Siege, L. and Ademasu, M. 2014. Crocodile management in Ethiopia. 60 pp. Shirley, M.H., Villanova, V.L., Vliet, K.A. and Austin, J.D. 2015. Genetic barcoding facilitates captive and wild
32 Crocodylus niloticus
management of three cryptic African crocodile species complexes. Animal Conservation, 18: 322–330. Spawls, S., Howell, K., Hinkel, H. and Menegon, M. 2018. Field guide to East African reptiles. First edit. Bloomsbury Publishing, New York, United States. 544 pp. The Government of Malawi 1994. National Park and Wildlife (Game Species) (Classification) Notice. Malawi Government. 259 pp. Thorbjarnarson, J.B., Messel, H., King, F.W. and Ross, P. 1992. Crocodiles: an action plan for their conservation. IUCN, Gland, Switzerland. 146 pp. Trutnau, L. and Sommerlad, R. 2006. Crocodilians. Their natural history and captive husbandry. Edition Chimaira, Frankfurt am Main, Germany. 646 pp. Uetz, P., Freed, P. and Hošek, J. 2019. The Reptile Database. http://www.reptile-database.org/ [Accessed 1/6/2019]. UNEP-WCMC and IUCN (2019), Protected Planet. Cambridge, UK: UNEP-WCMC and IUCN. Available at: www.protectedplanet.net. [Accessed 7/6/2019] Wallace, K.M., Leslie, A.J., Coulson, T. and Wallace, A.S. 2013. Population size and structure of the Nile crocodile Crocodylus niloticus in the lower Zambezi valley. Oryx, 47(3): 457–465. Wamisho, B. L., Bates, J., Tomkins, M., Islam, R., Nyamulania, N., Ngulube, C. and Mkandawire, N.C. 2009. Ward Round-Crocodile bites in Malawi: microbiology and surgical management. Malawi Medical Journal, 21(1): 29-31. Webb, G. 2019. Grahame Webb, Chair of IUCN SSC Crocodile Specialist Group. pers. comm. to UNEP-WCMC, 23/5/2019. Zisadza-Gandiwa, P., Gandiwa, E., Jakarasi, J., Westhuizen, H. van der and Muvengwi, J. 2013. Short Communication: Abundance, distribution and population trends of Nile crocodile (Crocodylus niloticus) in Gonarezhou National Park, Zimbabwe. Water SA, 39(1): 165–170.
33 Sphyrna lewini
CARCHARHINIFORMES: SPHYRNIDAE Sphyrna lewini II/B
SYNONYMS: Cestracion leeuwenii Day, 1865; Cestracion oceánica Garman, 1913; Sphyrna diplana Springer, 1941; Zygaena erythraea Klunzinger, 1871
COMMON NAMES: Scalloped hammerhead (EN), Requin marteau (FR), Cornuda común (ES)
RANGE STATES: Circumglobal distribution in coastal warm temperate and tropical seas between 46°N and 36°S
UNDER REVIEW: Kenya
EU DECISIONS: No current suspensions or opinions in place
IUCN: Endangered
Trade patterns
Sphyrna lewini was listed in CITES Appendix II on 12/06/2013 and in Annex B of the European Union (EU) Wildlife Trade Regulations on 10/08/2013, with a delayed entry into effect of 18 months (i.e. until 14/09/2014).
Kenya has submitted all annual reports 2014-2017, and has never issued an export quota for S. lewini. According to the CITES Trade Database, there were no direct exports of S. lewini from Kenya to either the EU- 28 or the rest of the world 2014-2016. Direct exports to the EU-28 in 2017 comprised low levels of live, wild- sourced individuals exported for commercial purposes to France and the Netherlands (seven individuals according to data reported by Kenya, two individuals according to data reported by importers); direct exports to the rest of the world comprised low levels of live, wild-sourced sharks for educational and commercial purposes (Table 1). No indirect trade in S. lewini originating in Kenya was reported to the EU-28 2014-2017.
Table 1: Direct exports of Sphyrna lewini from Kenya to the EU-28 (EU) and the rest of the world (RoW), 2014- 2017. All trade was in live, wild-sourced individuals reported by number. Kenya has submitted all annual reports 2014-2017. Importer Purpose Reported by 2014 2015 2016 2017 Total EU T Importer 2 2 Exporter 7 7 RoW E Importer 2 2 Exporter T Importer Exporter 3 3 Source: CITES Trade Database, UNEP-WCMC, Cambridge, UK, downloaded on 20/05/2019
There are no FAO global capture production data for Kenya for either the category ‘hammerheads, nei’ (not elsewhere included) or ‘scalloped hammerhead’. Sub-regional trade of shark products from Kenya to Somalia and vice versa was noted to continue undocumented (Rice, 2017).
34 Sphyrna lewini
Conservation status
Biology and distribution
Sphyrna lewini (scalloped hammerhead) is a large hammerhead shark, thought to be the most abundant species of the Sphyrnidae (Ebert and Stehmann, 2013). It is a coastal-pelagic species that has a circumglobal distribution in coastal warm temperate and tropical seas between 46°N and 36°S (Compagno, 1984). The species occurs over continental and insular shelves and in adjacent deep waters, from intertidal and surface waters to depths of 275 m, and has been observed entering enclosed bays and estuaries (Compagno, 1984). Some adult populations are known to form large aggregations at sea mounts (Ebert and Stehmann, 2013).
S. lewini is a highly mobile species (Compagno, 1984), but molecular studies have shown that males are much more mobile than females (Daly-Engel et al., 2012). Male S. lewini do not show any genetic population differences either between or within ocean basins, and as such are thought to travel long distances and facilitate gene flow across oceanic expanses (Daly-Engel et al., 2012). The frequency of these migrations, however, is unknown (Daly-Engel et al., 2012). In contrast, female S. lewini are thought to show site fidelity to single coastlines, archipelagos, or individual nursery areas (Daly-Engel et al., 2012). At least five genetically distinct populations of S. lewini have been identified: Northwest Atlantic, Caribbean Sea, Southwest Atlantic, Eastern Atlantic, and Indo-West Pacific (Chapman and Shivji, Nova Southeastern University unpublished data in Baum et al., 2007b). NOAA recognised six distinct population segments for S. lewini in 2013: Northwest Atlantic and Gulf of Mexico; Central and Southwest Atlantic; Indo-West Pacific; Central Pacific; and Eastern Pacific (NOAA, 2013; Figure 1). Within the Indo-West Pacific, stock structure was noted to be poorly understood (Rice, 2017).
Figure 1: Summary of S. lewini distinct population segment boundaries as reported by NOAA, 2013 (reproduced with permission)
As well as these sex differences, S. lewini show different levels of mobility depending on whether they are juveniles or adults. Juvenile S. lewini are found in coastal habitats, where they remain resident for a number of years before moving offshore as they grow (Compagno, 1984; Hoyos-Padilla et al., 2014). Natural predation of juveniles (by other carcharhinids as well as S. lewini adults) was reported to be high (Baum et al., 2007b).
35 Sphyrna lewini
S. lewini is viviparous (Compagno, 1984), with estimates of the length of gestation ranging from eight to 12 months (Hazin et al., 2001; White et al., 2008). Information on the reproductive periodicity of the species is conflicting, with some studies reporting that S. lewini females give birth every year (Cortés et al., 2010), and others estimating that females give birth every two years (Hazin et al., 2001; White et al., 2008). Other life history data for the species (including those needed to assess its intrinsic biological vulnerability such as the age and size at which individuals reach maturity, the average litter size, and the species’ natural mortality rate), vary depending upon the population sampled. Table 2 provides a summary of the range of life history parameters that have been estimated from populations in waters most relevant to this review (in this case, the Indo-West Pacific). In general, S. lewini is a long-lived, relatively slow-growing, and slow-reproducing species (Piercy et al., 2007; White et al., 2008), parameters which make it intrinsically vulnerable to overexploitation (Maguire et al., 2006; Baum et al., 2007b; White et al., 2008). A number of sources have placed the species within the FAO’s low productivity category (<0.14/yr) (CoP16 Prop. 43; Cortes et al., 2015).
Table 2: Summary of life history parameters for S. lewini. Life history parameters for the species vary upon the population sampled; this table shows figures calculated from samples of the Indo-West Pacific subpopulation. TL = total length, F = female, M = male. Life history parameter Value Location Source Size at maturity 210 cm TL F Northeastern Taiwan PoC Chen et al., 1988 in: Hazin 198 cm TL M et al., 2001 228.5 cm TL F Indonesia White et al., 2008 175.6 cm TL M 200 cm TL F Northern Australia Stevens and Lyle, 1989 140-160 cm TL M Age at maturity 13.2 years F Taiwan, PoC Chen et al., 1990 3.8 years M
(Assumes biannual growth band deposition7) 8.9 years M Indonesia Drew et al., 2015 13.2 years F
(Assumes annual growth band deposition) 5.7 years M Eastern Australia Harry, 2011 (no estimate for F)
(Assumes annual growth band deposition) Observed longevity 14 years F (331 cm TL) Northeastern Taiwan PoC Chen et al., 1990 10.6 years M (301 cm TL)
(Assumes biannual growth band deposition) No estimate for females Eastern Australia Harry, 2011 21 years M
(Assumes annual growth band deposition) 35 years F Indonesia Drew et al. 2015 No estimate for males
7 The most commonly used method to calculate a shark’s age is to look at band pairs in an individual’s vertebrae. While some studies assume that S. lewini puts down two band pairs per year (e.g. Chen et al., 1990), the majority of studies calculate growth rates on the assumption that one band is put down per year (Harry, 2011; Drew et al., 2015). Assuming annual ring growth rather than biannual ring growth results in (a) slower growth estimates and (b) higher estimates for the species’ age at maturity.
36 Sphyrna lewini
Life history parameter Value Location Source
(Assumes annual growth band deposition) Litter size 14-41 (mean = 25) Indonesia White et al., 2008 12-38 Northeastern Taiwan PoC Chen et al., 1988 in: Hazin et al., 2001 13-23 Northern Australia Stevens and Lyle, 1989
The natural mortality rate (using Jensen (1996)’s formula of Natural Mortality = 1.6k, where k is the von Bertalanffy growth completion rate) of the Indo-West Pacific subpopulation of S. lewini has been estimated at 0.126/year off the coast of Australia (Harry, 2011). Chen and Yuan (2006) calculated the species’ intrinsic rate of population increase in the Gulf of Mexico to be 0.086/year. Estimates of this parameter from Taiwan (Province of China), based on two band pairs per year, yielded a higher rate of 0.205/year (Liu and Tsai, 2011). No global estimates for the global population size, or the population size of the Indo-West pacific subpopulation, could be located, whether historical or current.
Conservation status
The global population of S. lewini was classified as Endangered, with an unknown population trend, in a 2007 IUCN assessment that is annotated as needing updating (Baum et al., 2007b). The assessment was made on the basis of major declines reported in many areas of the species' range, increased targeting for its high value fins, the species’ low resilience to exploitation, and largely unregulated, continuing fishing pressure from both inshore and offshore fisheries (Baum et al., 2007b). The Western Indian Ocean subpopulation was also assessed as Endangered, with a decreasing population trend, on the basis of continued high fishing pressure in the region and observed and inferred declines (Baum et al., 2007a). Overfishing was considered to be the principal threat both at the global and Western Indian Ocean sub-population level.
Fishery
S. lewini co-exists with other high-value pelagic species (FAO, 2010), and is taken as both catch and bycatch within Exclusive Economic Zones (EEZs) as well as on the high seas. The species is known to be susceptible to multiple different fishing gears including trawls and purse-seines (Hayes et al., 2009; Hazin et al., 2001; Baum et al., 2007b); however, it is generally considered to be most susceptible to gillnets (IOTC, 2018) and fixed bottom and pelagic longlines (Queiroz et al., 2016). Measures of susceptibility are a product of multiple factors including how likely a stock is to be encountered by a fishing fleet and how likely it is to be captured by the fishing gear (Benítez et al., 2015); and as such they may vary across different areas. The susceptibility of S. lewini to different fisheries, for example, changes as they mature; many juveniles are caught in coastal artisanal shark fisheries which tend to concentrate on nursery areas, whereas larger adults are caught as bycatch in pelagic fisheries for tuna and swordfish (Castillo-Géniz et al., 1998; Ruiz Alvarado and Mijangos López, 1998; White et al., 2008; Ebert and Stehmann, 2013). Sex disequilibrium in catches has also been observed, probably as a result of the species’ complex spatial dynamics (Branstetter, 1987; Hazin et al., 2001; Tavares and Arocha, 2008). S. lewini’s schooling habit makes the species vulnerable to capture in large numbers; and can make the species appear to be more abundant in landings data (Baum et al., 2007b).
Population trends
Species-specific landings and catch-per-unit effort (CPUE) data (used to calculate stock status and population trends) for elasmobranch taxa are generally scarce, principally because (a) until recently, shark catches have tended not to be reported to species level and (b) sharks are commonly caught as bycatch, and species that are caught as bycatch tend to be underreported. Hammerhead sharks are not thought to be commonly caught by the primary fisheries in the Indian Ocean (pelagic gillnet, pelagic longline and industrial purse seine); but in coastal fisheries within national waters, where the group is caught more frequently, little data collection
37 Sphyrna lewini occurs (Rice, 2017). Studies of S. lewini in the Indian Ocean have been noted to be particularly scarce (Miller et al., 2014; Rice, 2017), with the most recent Indian Ocean Tuna Commission (IOTC) stock assessment noting that the current paucity of information is not expected to improve in the short to medium term (IOTC, 2018). A 2016 workshop for IOTC countries noted that the lack of data for the region was such that the completion of NDFs for shark taxa “is all but impossible” (Rice, 2017).
In Oman, S. lewini was noted to have been displaced by smaller shark species during a two year study of landings conducted 2002-2003; however, the study was not able include measures of fishing effort (Henderson et al., 2007). An analysis of the Tanjung Luar artisanal shark fishery off East Lombok, Indonesia, found that the percentage catch accounted for by S. lewini decreased from 15% in 2001 to 2% in 2011 (FAO, 2013). Decreases in the average size of S. lewini individuals caught, which can indicate size truncation and overexploitation, have additionally been reported in Cochin Fisheries Harbour, India (CoP16 Prop. 43) and Taiwan, PoC (Joung et al., 2013 in: Miller, 2014).
In South Africa, CPUE data from protective gillnets off KwaZulu-Natal estimated that S. lewini had undergone a decline of 64% between 1978 and 2003 (Dudley and Simpfendorfer, 2006). Analyses of earlier data from shark nets in Main Beach and Brighton Beach (both near Durban, South Africa) estimated declines for S. lewini of 99.3% between 1952 and 1972, and 86% from 1961-1972, respectively (Ferretti et al., 2010).
In 2006, the FAO reported that the probable status of the shark and ray fishery in the southwestern Indian Ocean (where they were targeted predominantly by the Maldives, but were also targeted by Kenya, Mauritius, the Seychelles, South Africa and the United Republic of Tanzania) was fully exploited to overexploited (De Young, 2006). On the basis that no quantitative stock assessments or basic fishery indicators are available, the stock status and maximum sustainable yield (MSY) for S. lewini in the Indian Ocean are currently considered to be unknown (IOTC, 2018), with available evidence indicating “considerable risk to the stocks at current effort levels” (IOTC, 2014b). The IOTC Scientific Committee recommended that the IOTC Commission take a cautious approach by implementing some management actions for S. lewini, and to encourage its members to comply with their recording requirements (IOTC, 2018). The RFMO’s latest ecological risk assessment, consisting of a semi-quantitative analysis combining the biological productivity of the species and its susceptibility to each fishing gear type, gave S. lewini a low vulnerability ranking for longline gear (because, even though it was estimated to be one of the least productive shark species, it was also characterised by a lower susceptibility to longline gear) and also gave the species a low vulnerability ranking for purse seine gear (IOTC, 2018).
Illegal, unreported and unregulated (IUU) fishing of S. lewini has been noted to occur across its range, including the western Indian Ocean (Baum et al., 2007a). At an IOTC/CITES shark data mining workshop8 in 2016, it was reported that trade in CITES-listed shark species was known to be occurring, but that CITES Authorities were often unaware of it and were often unable to monitor trade or control it (Rice, 2017). The formal and informal markets for shark products were noted to be often indistinct; for example, shark products on sale or exported from Kenya were claimed in some cases to originate in Somalian waters (Rice, 2017), indicting cross-border movement without CITES permits. Traceability was also noted to be generally low (Rice, 2017).
Kenya
S. lewini is thought to have a near-coast wide range in Kenya (Kiilu and Ndegwa, 2013). No studies on the specific life history or biological parameters for the species’ Kenyan population could be located; the most
8 The IOTC/CITES Shark Data mining workshop was held in Victoria, Seychelles, in November 2016. Its main objectives were to conduct data mining to compile historical data for CITES-listed oceanic whitetip (Carcharhinus longimanus) and hammerhead sharks, and to develop descriptive indicators related to stock status. The workshop was led by a consultant (Dr Joel Rice) as part of the IOTC’s Working Party on Ecosystems and Bycatch (WPEB) Program of Work, and was funded by CITES. Prior to the workshop, IOTC contracting parties (CPCs) were asked to provide data to improve the status of information for populations of CITES-listed hammerhead sharks in the IOTC area. Data was submitted by Indonesia, Iran (Islamic Republic of), Seychelles, Kenya, the United Republic of Tanzania, Sri Lanka, Pakistan and Australia.
38 Sphyrna lewini relevant figures are detailed in Table 2, which shows data from studies sampling other locations in the Indo- West Pacific.
Sharks in the country are principally thought to be caught as by-catch by artisanal tuna fisheries, prawn trawls, sport fishing activities (Kiilu and Ndegwa, 2013) and longliners (Rice, 2017), though Kiilu and Ndegwa (2013) also reported that there were 490 fishers targeting sharks according to the Kenya Marine Frame Survey 2012. The majority of effort is assumed to be within the Kenyan EEZ (Rice, 2017). While shark meat is consumed domestically (including hammerhead meat, which is considered to be high quality), and was thought to be sourced from both domestic landings and imports (Vannuccini, 1999), fins were reported to be exported to the Asian market; particularly China and Hong Kong (Special Administrative Region of China) (Kiilu and Ndegwa, 2013). The shark fishery was considered to be important for both the economies and subsistence of local communities (Kiilu and Ndegwa, 2013), but the country’s small scale elasmobranch fishery was noted to have been decreasing since the early 1980s (Kiilu and Ndegwa, 2013). The fishery’s decreasing trend was described to have continued ‘alarmingly downward’ since 2000 (Kiilu and Ndegwa, 2013).
The Kenya Fisheries Service/Kenya Marine and Fisheries Research Institute estimated that the country had 414 artisanal vessels engaged in fishing for tuna and tuna-like species in coastal waters in 2017 (Mueni et al., 2018), whereas the entire artisanal fishing fleet was reported to consist of 3500 small scale crafts, usually used for single day fishing trips (Mueni et al., 2018). It was reported that the country’s artisanal fleet principally uses artisanal long line hooks, gillnets, monofilament nets and artisanal trolling lines (Mueni et al., 2018). Kenya’s other principal fishery (termed by the FAO as it’s EEZ fishery), mainly employs purse-seine and longline nets and targets tuna (FAO, 2016).
Reported landings of hammerheads in Kenya by the country’s artisanal fleet have fluctuated year-on-year 2014-2017 (Table 3); however, further historical data on hammerhead landings that were disaggregated from general shark landings could not be found.
Table 3: Catches of Sphyrnidae by the Kenyan artisanal fishing fleet according to data provided by Kenya as part of their national reports to the scientific committee of the IOTC, 2014-2017. Year 2014 2015 2016 2017 Catch (tons) 3.2 12.1 30.8 20.3
Both data from Kenya’s Department of State Fisheries and a study conducted by Kiilu and Ndegwa (2013) at six sampling points spanning the Kenyan coast from 2012-2013 have highlighted the importance of the Malindi- Uungwana ecosystem complex as a major shark fishing area in the country. The region accounted for 52% of shark biomass landed in Kenya according to Kenya’s State Department of Fisheries, and 98% of shark catches recorded by Kiilu and Ndegwa (2013). S. lewini and the closely related Sphyrna zygaena were the two most commonly landed species by the country’s small-scale and semi-industrial prawn trawl fleet in Uunwana Bay from 2012-2013, accounting for 30% and 27% of recorded catches during the study period, respectively (Kiilu and Ndegwa, 2013). The majority of individuals sampled from both the prawn trawl fleet and the artisanal fleet in Ungwana Bay were juveniles, with some neonate samples having fresh umbilical cords, indicating that Kenya’s principal shark fishing grounds consist of parturition and nursery sites for the species (Kiilu and Ndegwa, 2013). Most landings of S. lewini occured during the closed season for prawn trawl fishing (November to April), with a peak in January (Kiilu and Ndegwa, 2013). It is unclear if the prawn fisheries target other species during the closed season.
As a partially migratory marine species that is found in both coastal waters and the high seas, the management of S. lewini comes under the remit of policy frameworks operating on an international, regional, and national level. The family Sphyrnidae is listed under Annex I of the United Nations Convention for the Law of the Sea (UNCLOS), which is considered to be the principal framework convention for the management of the world’s oceans (United Nations General Assembly, 1982; Fischer et al., 2012). Under UNCLOS, coastal States and other States whose nationals fish in the region for listed species should cooperate directly or through appropriate international organizations with a view to ensuring conservation and promoting the objective of optimum
39 Sphyrna lewini utilization of such species throughout the region, both within and beyond the EEZ (United Nations General Assembly, 1982). Where no appropriate international organization exists, the coastal State and other States whose nationals harvest these species in the region area are asked to cooperate to establish such an organization and participate in its work (United Nations General Assembly, 1982).
S. lewini was also listed in Appendix II of CMS in 2014, and the species is covered by the CMS Memorandum of Understanding (MoU) on the Conservation of Migratory Sharks (CMS, 2016). Though non-binding, the MoU contains a conservation plan that aims to achieve and maintain a favourable conservation status for migratory sharks. Kenya is a signatory of the CMS Sharks MoU, but noted in its response to a survey by the CMS Secretariat on domestic legislation in 2016 that it did not prohibit shark finning, was not a member of a regional plan of action for sharks (RPOA-Sharks), and did not have any temporal or spatial closures for sharks or bans on the sale of shark fins/products (CMS Sharks MOS2 Inf. 17). A National Plan of Action for sharks has been in development since at least 2014 (Wekesa, 2014), and in 2018 was noted to have undergone an initial review by stakeholders (Mueni et al., 2018). No further information on the implementation of this plan was located.
IOTC measures: Kenya is a member of the Indian Ocean Tuna Commission (IOTC), a Regional Fisheries Management organisation that requires vessels operating within the IOTC area to undertake a number of measures regarding shark finning, bycatch, reporting, and data collection and research. The requirements of the IOTC, as well as the steps Kenya is known to have taken to fulfill them, are detailed below:
Finning: IOTC resolution 17/05 on the conservation of sharks caught in association with fisheries managed by IOTC states that Contracting Parties and Cooperating Non-Contracting Parties (CPCs) should take necessary measures to require their fishers to fully use any retained catches of sharks (IOTC, 2017). Full utilization is defined as retention by the fishing vessel of all parts of the shark except the head, guts and skins, to the point of first landing (IOTC, 2017). If sharks are landed fresh, CPCs should prohibit the removal of shark fins on board vessels (IOTC, 2017). If sharks are landed frozen, CPCs should require their vessels to not have on board fins that total more than 5% of the weight of sharks on board, up to the first point of landing (IOTC, 2017). CPCs that currently do not require fins and carcasses to be offloaded together at the point of first landing should take the necessary measures to ensure compliance with the 5% ratio through certification, monitoring by an observer, or other appropriate measures (IOTC, 2017).
In Kenya, shark carcasses caught by artisanal fisheries were reported to be fully utilised (Mueni et al., 2018), but the country noted in a response to a survey by the CMS Secretariat on domestic legislation in 2016 that it did not prohibit shark finning (CMS Sharks MOS2 Inf. 17).
Bycatch: The release of live sharks (especially juveniles and pregnant females) that are caught incidentally and are not used for food/subsistence is encouraged (IOTC, 2017). However, it is important to note that S. lewini has a very high at-vessel mortality rate: over 91% according to a study of US bottom longline vessels (70% young, 95.2% juvenile, 90.9% adult, N=455 individuals) (Morgan and Burgess, 2007), 62.9% according to Gulak et al. (2015)’s study of bottom longlines in the same region, and 100% in bottom-longline fishing experiments off Brazil (Afonso et al., 2011 in: Gulak et al., 2015).
Kenyan recreational trolling line fisheries were reported to have a voluntary shark release policy (Wekesa, 2014); however, no estimates for the at-vessel mortality rate of S. lewini caught on trolling lines could be located. Landings of hammerhead sharks from recreational trolling lines were reported to be very low, consisting of less than four individuals per year 2009-2014 (Wekesa, 2014; Ndegwa, 2015). It is unclear whether other Kenyan fisheries have release policies in place.
Reporting: IOTC members are required to submit annual reports for shark catches, including all available historical data, estimates and life status of discards (dead or alive), and size frequencies (IOTC,
40 Sphyrna lewini
2017). A lack of full compliance with IOTC data reporting measures for sharks was noted to be a general issue at a regional (Indian Ocean) level (Rice, 2017).
Data collection and research: Resolution 17/05 calls on IOTC members to undertake research to a) identify ways to make fishing gears more selective, b) improve knowledge on key biological/ecological parameters, life-history and behavioural traits, migration patterns of key shark species; c) identify key shark mating, pupping and nursery areas; and d) improve handling practices for live sharks to maximise post-release survival (IOTC, 2017). The IOTC Scientific Committee should annually review the information provided by CPCs, and, as necessary, provide recommendations to the Commission on ways to strengthen the conservation and management of sharks within IOTC fisheries (IOTC, 2017).
Kenya reported that it has planned numerous surveys to collect baseline data on catch composition and gear use by the country’s artisanal tuna fishery and small scale purse seine fishery (Mueni et al., 2018).
In 2014, the European Union submitted a proposal to the IOTC for a Resolution on a scientific and management framework on the conservation of shark species and on the protection of hammerhead sharks (family Sphyrnidae) caught in association with fisheries managed by IOTC (IOTC, 2014a). Taking into account the vulnerability of Sphyrna spp. and the indications of likely stock depletion in the Indian Ocean, the proposal contained provisional measures to prohibit the retention on-board of all species of the family Sphyrnidae, as well as the retention, transhipping, landing, storing, selling or offering for sale any part or whole carcasses of these species (IOTC, 2014a). The IOTC Commission considered the proposal at its eighteenth session, but agreement could not be reached and the proposal was deferred until the next meeting of the Commission (IOTC, 2014b). No evidence that the proposal has since been discussed or adopted could be found. There does not appear to be any regional cooperation to manage shared stocks (Rice, 2017), aside from measures in place as a result of membership of the IOTC.
Kiilu and Ndegwa (2013) reported that Kenya had no national or trans-boundary management systems in place for chondrichthyan populations in the country, but subsequently, in 2016, the FAO reported that the country had a “suite of management objectives and programmes, which include… Protection of endangered marine species from fishing activities such as… vulnerable shark species” (FAO, 2016). It is not clear what these measures are, and it was also noted that Kenya’s shrimp fishery is currently the only fishery with a management plan (FAO, 2016). Kenya’s 2018 National Report to the IOTC Scientific Committee noted that requirements under IOTC Resolution 13/06 on the conservation of oceanic whitetip sharks were being addressed through the 2016 Fisheries Management and Development Act (Government of Kenya, 2016); however, while the act itself provides for the declaration of endangered species of fish, the establishment of protected areas, gear restrictions, and the limitation of fishing activities, no shark-specific restrictions are included.
In the final summary report on the status of hammerhead sharks based on the results of the IOTC/CITES Shark Data Mining Workshop of 2016, Rice (2017) noted that increased observer monitoring would be vital to understanding the status of these species in the IOTC area. All Kenyan flagged and licenced foreign vessels are registered on the country’s vessel monitoring system (VMS), but the country’s artisanal tuna fishing vessels were noted to be too small to be fitted with VMS equipment (Mueni et al., 2018). Observer coverage of Kenyan flagged longliners was reported to be 100%, but small and medium tuna vessels were noted to be too small to accommodate observers (Mueni et al., 2018). A lack of capacity to monitor activities is thought to be the principal cause of IUU fishing in Kenya’s distant waters (FAO, 2016). References
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44 Appendix
Appendix Table 1: Purpose of trade Code Description
B Breeding in captivity or artificial propagation
E Educational
G Botanical garden
H Hunting trophies
L Law enforcement / judicial / forensic
M Medical (including bio-medical research)
N Reintroduction or introduction into the wild
P Personal
Q Circus and travelling exhibitions
S Scientific
T Commercial
Z Zoos
Table 2: Source of specimens Code Description
W Specimens taken from the wild
R Specimens originating from a ranching operation
D Annex A animals bred in captivity for commercial purposes and Annex A plants artificially propagated for commercial purposes in accordance with Chapter XIII of Regulation (EC) No 865/2006, as well as parts and
derivatives thereof
A Annex A plants artificially propagated for non-commercial purposes and Annexes B and C plants artificially propagated in accordance with Chapter XIII of Regulation (EC) No 865/2006, as well as parts and
derivatives thereof
C Annex A animals bred in captivity for non-commercial purposes and Annexes B and C animals bred in captivity in accordance with Chapter XIII of Regulation (EC) No 865/2006, as well as parts and derivatives thereof
F Animals born in captivity, but for which the criteria of Chapter XIII of Regulation (EC) No 865/2006 are not met, as well as parts and derivatives thereof
I Confiscated or seized specimens (to be used only in conjunction with another source code)
O Pre-Convention (to be used only in conjunction with another source code)
U Source unknown (must be justified)
X Specimens taken in “the marine environment not under the jurisdiction of any State”
45