Review of species selected from the Analysis of 2004 EC ...

Review of species selected from the Analysis of 2004 EC Annual Report to CITES

(Version edited for public release)

Prepared for the

European Commission Directorate General E - Environment

ENV.E.2. – Development and Environment

by the

United Nations Environment Programme World Conservation Monitoring Centre

August, 2007

Prepared and produced by: UNEP World Conservation Monitoring Centre, Cambridge, UK ABOUT UNEP WORLD CONSERVATION MONITORING CENTRE www.unep-wcmc.org The UNEP World Conservation Monitoring Centre is the biodiversity assessment and policy implementation arm of the United Nations Environment Programme (UNEP), the world’s foremost intergovernmental environmental organisation. UNEP-WCMC aims to help decision-makers recognize the value of biodiversity to people everywhere, and to apply this knowledge to all that they do. The Centre’s challenge is to transform complex data into policy-relevant information, to build tools and systems for analysis and integration, and to support the needs of nations and the international community as they engage in joint programmes of action. UNEP-WCMC provides objective, scientifically rigorous products and services that include ecosystem assessments, support for implementation of environmental agreements, regional and global biodiversity information, research on threats and impacts, and development of future scenarios for the living world. The contents of this report do not necessarily reflect the views or policies of UNEP or contributory organisations. The designations employed and the presentations do not imply the expressions of any opinion whatsoever on the part of UNEP, the European Commission or contributory organisations concerning the legal status of any country, territory, city or area or its authority, or concerning the delimitation of its frontiers or boundaries.

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Table of Contents 1. Introduction to the Analysis of the European Community Annual Report to CITES .......................... 2

2. Introduction to the species sheets ....................................................................................................... 2

SPECIES: Prionailurus bengalensis ................................................................................................... 7

SPECIES: Alligator mississippiensis ................................................................................................ 30

SPECIES: Caiman crocodilus crocodilus .......................................................................................... 45

SPECIES: Varanus salvator ............................................................................................................. 64

SPECIES: Euphorbia antisyphilitica ................................................................................................ 77

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1. Introduction to the Analysis of the European Community Annual Report to CITES

UNEP-WCMC undertakes an annual Analysis of the European Community Annual Report to CITES. This Analysis examines patterns of trade into the European Community, trade in groups of particular note, possible transgressions of suspensions and negative opinions, exports of native species, etc. The Analysis of the 2004 annual reports identified species for which:

• Imports were considered to be high volume according to thresholds which were determined by taxonomic group and CITES Appendix (Table 1).

• Imports of wild specimens had increased sharply in 2004, compared with imports between 1995-2004. Species were taken to qualify on the basis of the ‘sharp increase’ criterion if the volume of importer-reported wild-sourced imports during 2004 exceeded the upper 95% ‘confidence interval’ (i.e. the mean plus 1.96 standard deviations of the mean) for the average trade volume during 1995–2004. In such cases, the 2004 trade level was taken to show an increase in excess of what might be expected under ‘normal’ inter-annual fluctuations.

• Imports of wild specimens generally increased or decreased in 2002-2004, compared with the trade reported over the period 1995-2001. The overall increase criterion took into account more general trends over the period 1995–2004. This criterion compared the average volume of wild-sourced imports during 2002–2004 with that for preceding years, to highlight species that had experienced a significant increase or decrease over the ten-year period.

Table 1. Minimum level of trade required for selection on the basis of high trade volume.

Taxonomic group CITES Appendix

I II III

Mammals 100 10000 50000

Birds 100 10000 50000

Reptiles 100 50000 100000

Amphibians 100 50000 —

Fish 100 50000 —

Invertebrates 500 50000 100000

Plants (non-timber) 500 50000 100000

Plants (timber) — 1000 m3 5000 m3

All species initially identified by the criteria were subsequently inspected manually in order to exclude any for which the averages were disproportionately influenced by one or two atypical years.

2. Introduction to the species sheets

On the basis of this analysis, 49 Appendix II taxa were initially considered as candidates for review (Table 2). Following discussion with the European Commission, and considering previous SRG and CITES discussions, 13 species were selected for in-depth review. The following approach was taken: SRG 40: six in-depth species reviews and one summary review (for Python regius) were prepared;

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SRG 41: seven in-depth species reviews were prepared. Five are presented in this report and two (Python reticulatus and Scleractinia spp.) have been considered in another SRG 41 document. Additionally Cyclamen coum was selected for review on the basis of the increase in the 2007 CITES export quota for Turkey. Species were selected on the basis of particular patterns of trade in 2004. However, data for 2005 were available at the time of writing, and so were included in the species sheets. 2006 data were also summarised where available. Trade data in this report were downloaded on 30/03/2007. In the trade tables, “main” imports are summarised. These exclude trade terms involving a very small volume of trade (e.g. specimens, teeth). “Direct” exports are those where the origin and exporter are the same country. “Indirect” exports are those where the origin and the exporter are different countries. The table for direct exports to countries other than EU-27, may include trade included in the table for indirect exports to EU-27. Trade between EU-27 members is not included although some trade was reported from candidate countries to the EU during the pre-accession period of some Member States. The five species presented in this review were selected on the basis of the high volume of trade to the European Union in 2004. Alligator mississippiensis also showed a sharp increase in trade in 2004, and trade in Euphorbia antisyphilitica generally increased 2002-2004, compared with the trade reported over the period 1995-2001. Status information focuses on the main trading partners, with less detail provided for countries that are not involved in international trade.

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Table 2. Species considered for in-depth review

Species EC Annex IUCN

Red List Criteria met Notes Review?

Monodon monoceros A/B DD Sharp increase Under discussion in CITES arena. No

Pseudalopex griseus B LC High volume,

overall increase

Reviewed in 2000 for SRG18 (positive opinion formed for Argentina and Chile), but EC imports of wild-sourced skins have shown a notable increase since then.

SRG 40

Lynx canadensis B LC Overall increase

Reviewed in 2001 for SRG20 (positive opinion formed for Canada and the USA), but EC imports of wild-sourced skins have increased since. It was felt that a review would be useful when considering the proposal on Lynx rufus submitted for discussion at CoP 14.

SRG 40

Lynx rufus B LC High volume,

overall increase

Reviewed in 2000 for SRG18 (positive opinion formed for Canada, Mexico and the USA), but EC imports of wild-sourced skins have shown a substantial increase since. It was felt that a review would be useful when considering the proposal on Lynx rufus submitted for discussion at CoP 14.

SRG 40

Panthera leo B VU Overall decrease Included in 2004 analysis on basis of an overall decrease in EC imports of wild-sourced trophies.

No

Prionailurus bengalensis

B LC High volume

Brief review of trade from Macao in 1999 for SRG11, but EC imports of wild-sourced skins / plates remain (variable but) relatively high.

SRG 41

Arctocephalus pusillus

B LC High volume,

overall increase

Reviewed in 1997 for SRG2 (positive opinion formed for all countries of origin), but EC imports of wild-sourced skins have shown a notable increase since.

SRG 40

Pecari tajacu B LC High volume

Exports from Peru included in the CITES Significant Trade Process in 2001, and skin exports from Peru have since been below level set by export quota.

No

Tayassu pecari B LC High volume,

overall decrease

Exports from Peru included in the CITES Significant Trade Process in 2001, and skin exports from Peru have since been below level set by export quota.

No

Agapornis roseicollis B LC High volume Veterinary ban on EC imports of wild-sourced birds.

No

Amazona aestiva B LC Overall increase Veterinary ban on EC imports of wild-sourced birds. No

Aratinga mitrata B LC Sharp increase, overall increase

Veterinary ban on EC imports of wild-sourced birds.

No

Cacatua ducorpsii B LC Sharp increase, overall increase

Veterinary ban on EC imports of wild-sourced birds.

No

Cyanoliseus patagonus

B LC Overall decrease Veterinary ban on EC imports of wild-sourced birds.

No

Myiopsitta monachus

B LC Overall decrease Veterinary ban on EC imports of wild-sourced birds.

No

Poicephalus senegalus

B LC High volume Veterinary ban on EC imports of wild-sourced birds.

No

Psittacus erithacus B LC High volume Veterinary ban on EC imports of wild-sourced birds.

No

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Leiothrix lutea B LC High volume Veterinary ban on EC imports of wild-sourced birds.

No

Gracula religiosa B LC Overall decrease Veterinary ban on EC imports of wild-sourced birds.

No

Indotestudo elongata B EN Overall decrease Included in 2004 analysis on basis of an overall decrease in EC wild-sourced imports.

No

Alligator mississippiensis

B LC High volume, sharp increase

Reviewed in 2001 for SRG20 (positive opinion formed for the USA), but EC imports of wild-sourced skins showed a sharp increase in 2004.

SRG 41

Caiman crocodilus crocodilus

B NE High volume C. crocodilus reviewed in 1999 for SRG15, but status of subspecies crocodilus in Venezuela not specifically covered.

SRG 41

Caiman yacare B LC High volume,

overall increase

Steady increase in EC imports of wild-sourced skins in recent years, but positive opinion formed for Bolivia in 2006 Norway (recent opinion Bolivia)

No

Iguana iguana B NE High volume EC imports almost exclusively from captive-bred sources.

No

Tupinambis merianae

B NE High volume High volume of wild-sourced imports. SRG 40

Varanus salvator B NE High volume

Exports from Indonesia (the principal origin of EC imports) reviewed in 1999 for SRG13 and positive opinion formed. Volume of EC imports remains high, but is considerably below the export quota for Indonesia.

SRG 41

Python brongersmai B NE High volume Relatively high volume of imports into the EC, mainly from Malaysia.

SRG 40

Python regius B NE Overall decrease Recent reviews of exports from various exporting countries.

SRG 40 - summary

review

Python reticulatus B NE High volume Reviewed in 2002 for SRG24 (negative opinion formed for Malaysia).

SRG 41

Eunectes notaeus B NE Overall decrease Included in 2004 analysis on basis of an overall decrease in EC wild-sourced imports.

No

Ptyas mucosus B NE Overall decrease

Included in 2004 analysis on basis of an overall decrease in EC wild-sourced imports (owing to trade suspension for specimens not from registered stockpiles).

No

Hoplobatrachus tigerinus

B LC High volume EC imports mainly from captive-bred sources.

No

Polyodon spathula B VU Overall increase EC imports mainly of F1 captive-born specimens. Further positive opinion (USA) in 2001

No

Scleractinia spp. B NE High volume, sharp increase, overall increase

Trade discussed in 2004 analysis at higher-taxon level.

SRG 41

Acropora tenuis B NE High volume, sharp increase

Included in 2004 analysis on basis of high volume of imports for scientific purposes.

No

Agaricia agaricites B NE Sharp increase

Included in 2004 analysis on basis of a sharp increase in EC imports in 2004, but negative opinion subsequently formed for Haiti at SRG33 in 2005.

No

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Galanthus elwesii B NE High volume

CITES Significant Trade review of bulb trade from Turkey in 1999, and Turkey export quota has not increased significantly since then. 2007 export quota was the same as 2006 quota.

No

Galanthus woronowii

B NE High volume,

overall increase

CITES Significant Trade review of bulb trade from Turkey and Georgia in 1999, but Georgia export quota has almost doubled since then. Germany reported on this species at SRG20. 2007 export quota was the same as 2006 quota.

No

Cyathea spp. B NE/DD/NT/VU

Overall increase Trade discussed in 2004 analysis at higher-taxon level.

No

Euphorbia antisyphilitica

B NE High volume,

overall increase EC imports of wild-sourced wax have shown a notable increase in recent years.

SRG 41

Euphorbia primulifolia

B VU Sharp increase

Included in 2004 analysis on basis of sharp increase in EC imports in 2004, albeit still at a relatively low volume. [Globally threatened. Negative opinion (later removed) in 2006.]

No

Aloe ferox B NE High volume, sharp increase, overall increase

Review by TRAFFIC in 2006 for SRG37. Also excluded from further review in the Sig Trade process for South Africa at PC16.

No

Cyclamen cilicium B NE High volume

CITES Significant Trade review of bulb trade from Turkey in 1999, and Turkey export quota has not increased significantly since then. Review could be considered pending 2007 export quota.

No

Cyclamen coum B NE High volume,

overall increase

CITES Significant Trade review of bulb trade from Turkey in 1999, but Turkey export quota has doubled since then. Review could be considered pending 2007 export quota.

SRG 41

Cyclamen hederifolium

B NE High volume

CITES Significant Trade review of bulb trade from Turkey in 1999; Turkey export quota has increased by c.50% since then, but has not been fully utilised. Review could be considered pending 2007 export quota.

No

Pericopsis elata B EN High volume Exports from Cameroon reviewed in 2005 for SRG34

No

Swietenia macrophylla

B VU High volume Currently being discussed in the CITES arena and subject of numerous discussions by the SRG.

No

Prunus africana B VU High volume Reviewed in the CITES Significant Trade process.

No

Guaiacum sanctum B EN Sharp increase

Reviewed in 2003 for SRG26 (negative opinion formed for all countries of origin, with exceptions for certain stocks in Mexico).

No

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REVIEW OF SELECTED SPECIES FROM ANALYSIS OF EC ANNUAL REPORTS 2004

MAMMALIA FELIDAE

SPECIES: Prionailurus bengalensis

SYNONYMS: Felis bengalensis, Felis euptilura COMMON NAMES: Chinese Bengaalse kat (Dutch), Leopard Cat (English),

Chat de Chine (French), Chat-léopard de Chine (French), Chat-léopard du Bengale (French), Gato bengalí (Spanish), Gato de Bengala (Spanish), Gato leopardo chino (Spanish), bengalisk katt (Swedish), dvärgtiger (Swedish), dvärgtigerkatt (Swedish), kattamurkatt (Swedish), leopardkatt (Swedish)

RANGE STATES: Afghanistan, Bangladesh, Bhutan, Brunei

Darussalam, Cambodia, China, Democratic People's Republic of Korea Hong Kong, China, India, Indonesia, Japan, Korea, Republic of, Lao People's Democratic Republic, Macao, China Malaysia, Myanmar, Nepal, Pakistan, Philippines, Russian Federation, Singapore, Taiwan, Province of China, Thailand, Viet Nam

RANGE STATES UNDER REVIEW: All, with emphasis on China, the main exporter IUCN RED LIST: Least Concern (2001) PREVIOUS EC OPINIONS: A trade suspension has been in place for Macao,

China from 22/12/1997 until present (10/05/2006). TRADE PATTERNS: The Leopard Cat, Prionailurus bengalensis was selected for review following a high volume of trade of wild sourced skins and plates into the European Community over the period 1995-2004 (UNEP-WCMC, 2006).

China was the main exporter of the Leopard Cat between 2001-2005. In this time period, no trade was reported from Bangladesh, Bhutan, Brunei Darussalam, Democratic People's Republic of Korea, India, Indonesia, Nepal, Pakistan, Philippines, Macao, China and Hong Kong, China. No trade has been reported from Afghanistan between 2001-2005 except for the import of 20 skins (Source I) in 2004 reported by the United Arab Emirates. 2006 trade data available at the time of writing indicated that the UK reported the import of 100 Prionailurus bengalensis plates and 100 Prionailurus b. chinensis plates from China. San Marino reported the export of 250 Prionailurus bengalensis skins (Purpose T, Source W), origin China, to Italy. A total of 49 Prionailurus bengalensis garments (Purpose T, Source W), origin China were re-exported from Hong Kong, the United States and Switzerland to Spain (46), the UK (1) and Germany (2) (all reported by importer) respectively. 100 Prionailurus bengalensis skins (Purpose T, Source W) and 3 Prionailurus bengalensis garments, all orgin China, were re-exported (as reported by the exporter) to countries other than EU-27. In China commercial exploitation has been heavy, especially in the south-west (Nowell & Jackson 1996). Wang (1998) stated that the species was one of the main export products during the 1960s and 1970s when the volume of exports was suggested to be about 200,000-

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250,000 individuals per year. According to Wang (1998) statistics, the ”skin purchasing amount” in 1964 and 1965 was 222,000 and 172,000 skins respectively, from the Provinces of Yunnan, Guizhou, Guangxi, Hunan, Hubei and Jiangxi. In 1978 and 1979, this quantity decreased to 126,000 and 141,000, respectively. In the 1980s, the ”skin purchasing amount” of skins was around 150,000-200,000. Following the transfer of the Chinese population of P. b. bengalensis from CITES Appendix I to II in 1985, exports increased from 138,000 skins in 1984 to 200,000 skins in 1987-1988 (Wang, 1998). Hundreds of thousands of Leopard Cat skins were exported every year until Europe suspended imports in 1988 over concern for the species status (WCMC & IUCN/SSC Trade Specialist Group, 1992; Nowell & Jackson 1996). Huge stockpiles, estimated at 800,000 Leopard Cat skins, built up awaiting export in 1989 from Yunnan and Guizhou (Yu Jinping, 1991 in WCMC & IUCN/SSC Trade Specialist Group, 1992). By the 1990s, the number of traded skins was at about half the level exported during the 1960s (Wang, 1998). In Hong Kong, China in 2000, customs officers seized a Leopard Cat from a shipment which was being unloaded at the Tuen Mun Public Cargo Working Area, Hong Kong en route from the Philippines and Indonesia, to Shenzhen, China. (TRAFFIC, 2000). In Japan, the government has not issued any export permits since 1993, although ISEC (2001) reported that 80,000 to 150,000 skins were still exported to Russia. Japan is also a major cat fur consumer in Asia and Leopard Cat skins were considered a preferred choice, with the import of as many as 50,000 skins per year (ISEC, 2001). At least 1,128 Leopard Cat skins were exported from Thailand between 1967 and 1971 (Humphrey & Bain, 1990). Nilsson (2005) suggested that, in the 1970s, Leopard Cats were captured in Thailand, smuggled into Lao PDR, where they received Lao certificates of origin, and then re-exported via Bangkok airport. No quotas have been set for Prionailurus bengalensis in any country. Table 3. Main direct exports of Prionailurus bengalensis to EU-27, 2001-2005 excluding within EU trade.

Taxon Exporter Importer Term Source Purpose Reported by 2001 2002 2003 2004 2005 Total

P. bengalensis Canada Czech Rep. live C B Importer 2 2

Exporter

T Importer

Exporter 1 1

Z Importer

Exporter 2 2

Germany live C T Importer

Exporter 1 1

Hungary live C T Importer

Exporter 1 1

UK live C T Importer

Exporter 3 1 4

China France plates W T Importer 77 77

Exporter

Germany plates W T Importer 20 749 1391 215 50 2425

Exporter

skins W T Importer 900 900

Exporter

Italy plates W T Importer

Exporter 200 200


9


Exporter

Spain skins W T Importer 200 200 716 1116

Exporter

UK plates W T Importer 792 6986 450 750 8978

Exporter

skin

pieces W T Importer 2298 100 525 2923

Exporter

skins W T Importer 2877 4650 950 1103 9580

Exporter

small leather product

s W T Importer 200 260 460

Exporter

specime

ns W T Importer 204 204

Exporter

Japan UK bodies W E Importer

Exporter 1 1

Malaysia Bulgaria live W Z Importer 4 4

Exporter 4 4

Myanmar Czech Rep. live W T Importer 2 2

Exporter 2 2

USA UK live F T Importer

Exporter 1 1

P. b. chinensis China France garmen

ts W T Importer

Exporter 125 125

plates W T Importer 364 364

Exporter 519 164 200 77 960

skin

pieces W T Importer 500 500

Exporter


Exporter

Germany garmen

ts W T Importer

Exporter 50 50

plates W T Importer

Exporter 20 819 1421 175 2435

skins W T Importer

Exporter 900 900

Greece plates W T Importer

Exporter 130 60 190


Exporter

Italy garmen

ts W T Importer 20 8 28

Exporter 270 160 430

plates W T Importer 100 100

Exporter 14789 8004 7226 5300 1215 36534

skin W T Importer 330 330

10


pieces

Exporter

skins W T Importer 17382 14182 18807 6610 725 57706

Exporter 3100 5655 15340 300 24395

Spain plates W T Importer

Exporter 200 200 716 1116

UK plates W T Importer

Exporter 5110 5061 4100 2114 753 17138

plates

(pieces) W T Importer

Exporter 460 460

garmen

ts W T Importer

Exporter 38 520 450 1008

skins W T Importer

Exporter 6000 3000 9000

Table 4. Main indirect imports of Prionailurus bengalensis and subspecies P. b. bengalensis to EU-27, 2001-2005. All source W, Purpose T unless otherwise indicated.

Taxon Origin Exporter Importer Term Reported by 2001 2002 2003 2004 2005 Total

P. bengalensis China Egypt Germany skins Importer

1 (Purp.

P) 1

Exporter

1 (Purp.

P) 1

Hong Kong Germany plates Importer 300 300

Exporter

Spain garments Importer 24 39 63

Exporter

Japan Italy

small leather

products Importer

Exporter 7 7

Spain garments Importer 1 1

Exporter

large leather

products Importer 1 1

Exporter

small leather

products Importer

Exporter 2 2

San Marino France garments Importer 8 8

Exporter

Italy plates Importer

Exporter 300 300

USA France garments Importer 7 7

Exporter 3 3

skin pieces Importer

Exporter 4 4

Germany garments Importer 4 1 5

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Taxon Origin Exporter Importer Term Reported by 2001 2002 2003 2004 2005 Total

Exporter 1 1

Greece skin pieces Importer

Exporter 1 1

Italy garments Importer 5 2 1 8

Exporter 1 1

skin pieces Importer

Exporter 5 2 7

Unknown Hong Kong Slovenia skins Importer

1 (Purp.

E, Source

I) 1

Exporter

Russian Fed. United Kingdom garments Importer

1 (Source

I) 1

Exporter

P.b chinensis China Canada Italy skins Importer 150 150

Exporter 150 150

Hong Kong Greece plates Importer

Exporter 20 20

Italy garments Importer 85 17 102

Exporter

plates Importer

Exporter 5 5

skins Importer

Exporter 1998 1998

San Marino Italy skins Importer 200 200

Exporter

skins (kg) Importer 100 100

Exporter

Singapore Italy garments Importer 1 1

Exporter

Switzerland Italy garments Importer 2 2

Exporter

Unknown Various Greece skins Importer 250 250

Exporter

Table 5. Main direct exports of Prionailurus bengalensis to countries other than EU-27, 2001-2005.

Taxon Exporter Term Source Purpose Reported by 2001 2002 2003 2004 2005 Total

P.bengalensis Afghanistan skins I Importer 20 20

Exporter

Belgium live C B Importer 2 3 5

Exporter 5 2 7

P Importer 1 1 1 3

Exporter 1 1 1 3

T Importer 2 2 5 6 15

Exporter 10 6 16

F T Importer 2 2

Exporter

Cambodia specimens W S Importer 13 13

Exporter 28 28

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Canada live C B Importer

Exporter 1 1

P Importer 2 2 1 3 3 11

Exporter 1 4 4 9

T Importer 33 5 1 39

Exporter 17 7 2 26

F P Importer 1 1

Exporter

T Importer 4 1 5

Exporter

China garments I P Importer 2 2

Exporter

W T Importer 121 95 66 282

Exporter

hair R S Importer

Exporter 3 3

plates W T Importer 200 1910 1062 480 490 4142

Exporter

skin pieces R T Importer 1000 1000

Exporter

skins R S Importer

Exporter 4 4

W T Importer 3000 1304 500 4804

Exporter

small leather products W T Importer 60 60

Exporter

specimens R S Importer 4 4

Exporter 16 16

Czech Rep. live C B Importer

Exporter 3 2 5

P Importer 1 1

Exporter

T Importer 6 6

Exporter

F B Importer 3 3

Exporter

Germany live C Z Importer 3 3

Exporter 3 3

Malaysia live W S Importer

Exporter 1 1

Specimens (flasks) W S Importer 20 20

Exporter 20 20

Rep. of Korea

derivatives I Q Importer 1 1

Exporter

live C Z Importer 2 2

Exporter 2 2

Russian Fed. specimens C S Importer 6 6

13


Exporter

Singapore live C Z Importer 2 2 3 3 10

Exporter 2 2

Thailand live C B Importer 10 10

Exporter

specimens C S Importer 231 231

Exporter 231 231

USA live C T Importer

Exporter 7 1 1 1 10

F B Importer

Exporter 1 1

P Importer

Exporter 3 1 1 5

T Importer

Exporter 3 1 4

Unknown bodies I P Importer 1 1

Exporter

Viet Nam bodies I P Importer 2 2

Exporter

W P Importer 1 1

Exporter

trophies I P Importer 3 3

Exporter

W P Importer 2 1 3

Exporter P. b. bengalensis China skins I Importer 1 1

Exporter

Czech Rep. live C B Importer

Exporter 3 3

Taiwan, Prov. of China specimens C S Importer 2 2

Exporter

Thailand skins W P Importer 1 1

Exporter

Unknown bodies I P Importer 1 1

Exporter

trophies W P Importer 1 1

Exporter

Viet Nam bodies I P Importer 3 3

Exporter

W P Importer 1 1

Exporter

trophies W P Importer 2 2

Exporter P. b. chinensis China garments W P Importer 1 1

Exporter

T Importer

Exporter 123 114 72 309

plates W T Importer

14


Exporter 205 2300 1512 400 140 4557

plates (pieces) W T Importer

Exporter 25 25


Exporter 4000 2004 2500 8504

P. b. euptilura Czech Rep. live C B Importer

Exporter 2 2

T Importer

Exporter 4 4

Poland live C Z Importer 2 2

Exporter 1 1

Russian Fed. specimens C S Importer

Exporter 6 6 Prionailurus spp. Laos feet I P Importer 1 1

Exporter

Table 6. Main direct exports, from China to all countries, of wild P. bengalensis and subspecies.

Term Reported by 2001 2002 2003 2004 2005

garments Importer 1 121 95 86 8

Exporter 38 123 114 1037 610

plates Importer 1012 9645 3267 1545 477

Exporter 21043 16548 15305 8049 2185

skin pieces Importer 2798 330 100 525

Exporter

skins Importer 23459 17096 24673 7745 1828

Exporter 7100 14559 20840 300

small leather products Importer 60 200 260

Exporter

specimens Importer 204

Exporter

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CONSERVATION STATUS IN RANGE STATES: Considered to be one of the most common and the most widely distributed of Asian cats (Santiapillai & Supraham, 1985; Bain & Humphrey, 1996). Its geographical range extends from the dense tropical forests of Sumatra to the Manchurian and Siberian taiga, as far north as North Korea and the Amur basin and as far south as Bali, west towards Pakistan through northern India, the southern Himalaya, Bangladesh, Myanmar and east to the Indo-China region and it has colonised many islands (due to their ability to swim) in the Philippines, Indonesia and near Japan (Cat Survival Trust, 1996). A number of studies have indicated that the Leopard Cat is a generalist and opportunist predator (Rabinowitz, 1990; Grassman et al., 2005; Rajaratnam et al., 2007). It is highly adaptable to different climates and habitats (Santiapillai & Supraham, 1985) and may be present in scrublands, grasslands, secondary forest, tropical rainforest, temperate broadleaf, coniferous forests, semi-deserts, and agricultural regions (especially near water), at up to 3,000 m in parts of their range (Nowell & Jackson, 1996). They can be found close to villages and seemed to be relatively tolerant of human activity (Cat Survival Trust. 1996). The taxonomic placement of the Leopard Cat has varied over the years. The species was described in the genus Felis, and this was followed by many authors, including Honacki et al. (1982). However, Wozencraft (1993, in Wilson & Reeder, 1993) split the genus Felis into 13 genera and placed bengalensis in the genus Prionailurus, along with the Flat-headed Cat P. planiceps, Rusty-spotted Cat P. rubiginosus and Fishing Cat P. viverrinus. This classification was maintained by Wozencraft (2005, in Wilson & Reeder, 2005). Wozencraft (1993) treated the Iriomote Cat P. iriomotensis as a subspecies of P. bengalensis but, later (2005), he treated it as a separate species. Wilson & Reeder (2005) recognised 10 subspecies of P. bengalensis (see Table 8) Table 7. P. bengalensis subspecies.

Subspecies Region of distribution Appendix/Annex

P. b. alleni China (Hainan) Appendix II, Annex B

P. b. bengalensis

Bangladesh, Cambodia, China (Guizhou, Sichuan, Yunnan), India (south of the Himalayas), Lao People’s Democratic Republic, Myanmar, Pakistan, Thailand, Vietnam

Appendix I, Annex A (India, Thailand, Bangladesh) and Appendix II, Annex B for all other populations

P. b. borneoensis Brunei, Indonesia (Kalimantan), Malaysia (Sabah, Sarawak)

Appendix II, Annex B

P. b. chinensis China (Anhui, Fujian, Guangdong, Guangxi, Hong Kong, Hunan, Hubei, Jiangxi, Jiansu, Macao?, Taiwan, Zhejiang)


P. b. euptilura

China (northern provinces), Japan (Tsushima), Democratic People’s Republic of Korea, Republic of Korea, Russian Federation


P. b. heaneyi Philippines (Palawan) Appendix II, Annex B

P. b. horsfieldii Bhutan, India (Arunachal Pradesh, Jammu & Kashmir, Himachal Pradesh, Uttarakhand, Sikkim), Nepal


P. b. javanensis Indonesia (Bali, Java, Lombok) Appendix II, Annex B

P. b. sumatranus Indonesia (Sumatra), Malaysia (Peninsular Malaysia), Singapore


P. b. trevelyani Afghanistan, India (Himachal Pradesh, Jammu & Kashmir), Pakistan


P. b. rabori Philippines (Busuanga, Cebu, Negros, Panay)


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The nominate race of the Leopard Cat, P. b. bengalensis, was listed in CITES Appendix I in 1975, but this was revised in 1985 when the Chinese population of P. b. bengalensis was transferred to Appendix II and all other populations of the species were included in Appendix II. The situation was redefined again in 1995, when the Appendix I listing for P. b. bengalensis was restricted to the populations of Bangladesh, India and Thailand, leaving the remaining populations (Cambodia, China, Lao People’s Democratic Republic, Myanmar, Pakistan and Vietnam) in Appendix II.

The main threats to the Leopard Cat were reported to be habitat destruction and fragmentation, exploitation for trade including the the pet industry and pelts/fur industry, local consumption and hunting for meat, and the use of bones for medicinal purposes (Santiapillai & Supraham, 1985, WCMC & IUCN/SSC Trade Specialist Group, 1992; Ng & Wee, 1994; Nowell & Jackson, 1996; Duckworth, 1997; Wang, 1998; Jorio, 2000; ISEC, 2001; Indian Tiger Welfare Society, 2005; Srinivasulu, 2005; Alzan & Sharma, 2006). The harvest of wild animals to breed with domestic cats was also considered a threat to wild Leopard Cat populations (ISEC, 2001). Pelts of small spotted cats have been in high demand for clothing since the 1960s. More recently, high profile conservation measures on trading the fur of large cats, have shifted the emphasis of trading to the smaller wild cat species including the Leopard Cat which is under continued threat from hunting in many parts of its range (Indian Tiger Welfare Society, 2005). Nilsson (2005) suggested that following the decline in Latin American small cats, the Leopard Cat was overexploited and became endangered in India, Bangladesh and Thailand. Although some subspecies may be threatened, the species was considered sufficiently widespread to withstand a lot of human disturbance and exploitation (Cat Survival Trust. 1996). However, Cat Survival Trust (1996) indicated that there was insufficient information about the numbers of Leopard Cats in the wild to assess their status. Considered to be vulnerable by Gurung & Singh (1996). Nowell & Jackson (1996) reported that the species was protected at the national level in part of its range, with hunting prohibited in Bangladesh, Cambodia, Hong Kong (China), India, Indonesia, Japan, Malaysia (except Sabah), Myanmar, Nepal, Pakistan, Philippines, Russia, Thailand and Taiwan, and hunting and trade regulations are in place in the Republic of Korea, Laos and Singapore. AFGHANISTAN: Records are outlined in Hassinger (1973) and include a single skin purchased by the Third Danish Expedition in 1948-1949 and statements from Kullman (1965) that it was previously known from Afghanistan. However, the references quoted do not include any mention of the species. Its occurrence in monsoon forests of Afghanistan was also mentioned without details by Habibi (1977). The distribution of P. bengalensis is therefore debatable and the Leopard Cat is probably an occasional visitor rather than a permanent resident (Hassinger, 1973). P. b. trevelyani has been occasionally recorded in the forested regions of eastern Afghanistan: Kunar and Pech Valleys, Dare Nur (Anon., 1981a). Satellite imagery has shown that conifer forests in the provinces of Nangarhar, Kunar and Nuristan have shrunk by more than half since 1978. A study by UNEP said that hunting, mainly for meat and furs, was significantly reduced during the Soviet occupation, but has increased since then (the Leopard Cat was not specifically mentioned in this report) (UNEP, 2003). It is thus likely that carnivore species dependent on forest habitats have also been reduced. There is very little information on this species and it was not mentioned in the Wildlife Conservation Society’s Afghanistan Biodiversity Conservation Project (WCS, 2007). BANGLADESH: Occurrence reported in Bangladesh (Gurung & Singh, 1996) as P. b. bengalensis; its presence has been recorded in the tea plantations of Sylhet and Chittagong in the East and south-east of the country, where it was found in pet trade markets of Chittagong (Khan, 1985). Its distribution probably never extended to the Sunderbans and other mangrove formations (Khan, 1985). The Bangladesh Wild Life (Preservation) Order, 1973 deals with game and

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protected animals. Protected animals (including the Leopard Cat) cannot be hunted, killed or captured except for protecting life, crops or livestock (Ministry of Environment and Forests, 2003). P. bengalensis was considered ‘not common anywhere’ by Khan (1985). BHUTAN: Occurs in Bhutan (Gurung & Singh, 1996) as P. b. horsfieldii (Chakraborty, 1975; Ellerman & Morrison-Scott, 1966). P. b. horsfieldii was recorded in Bhutan in 1939; however, there seems to be a degree of doubt over whether the specimen originated in Upper Assam, India (Chakraborty, 1975). The Leopard Cat was listed as ‘Totally Protected’ in the Forest and Nature Conservation Act of Bhutan, 1995 and therefore may not be killed, injured, destroyed, captured, collected or otherwise taken except with a special permit or in self defence. BRUNEI DARUSSALAM: ‘Presumably occurs’ as P. b. borneoensis, which is present elsewhere in mainland Borneo (Medway, 1977). However, it was not recorded during a wildlife survey in 1984 (Bennett et al., 1984). The Wild Life Protection Act (s.1301s) did not list the Leopard Cat as a protected species but protected all animals within a wildlife sanctuary (Government of Brunei, 1981). CAMBODIA: The Leopard Cat is found in Cambodia as P. b. bengalensis (Osgood, 1932; Ellerman & Morrison-Scott, 1966) and was poerted to be the most commonly found species of cat (Duckworth et al., 2005). Cambodia supports by far the most extensive areas, in the Indochinese region, of deciduous dipterocarp forest habitat and holds the least affected populations of large vertebrates associated with these habitats (Duckworth et al., 2005). A survey undertaken by Duckworth et al. (2005) recorded >85 Leopard Cats during several periods of field work from 1997-2005. The Leopard Cat was also recorded in another study (in central Cambodia, December 1998) with an animal captured from the local area being held at the Phet Sanday Fisheries Office at the mouth of the Stung Sen (Goes & Hong, 2002 in Campbell et al., 2006).

CHINA: The Leopard Cat is found throughout the country except in Xinjiang, Inner Mongolia and Qinghai. Five subspecies are recognized: P. b. euptilura, in the northeast, north and northwest; P. b. chinensis, in the east, central and south; P. b. bengalensis in Yunnan, west Guizhou and west Guangxi; P. b. alleni, restricted to Hainan (Corbet & Hill, 1992). A recent study in Niubeiliang National Nature Reserve (NNR, 108°45’–109°04’ E, 33°47’–33°56’ N) in Shaanxi Province recorded the Leopard Cat at high altitudes with a vertical elevation range of about 1000 m and a maximum altitude of 2200 m (Zeng et al., 2007). Wang (1998) stated that the species was formerly common in China and was exported in high numbers in the 1960s and 1970s, when the total population was estimated at more than 1 million. Since the 1970s, the population size was reported to have declined greatly and in some places it had become very rare (particularly in north and east China). (Wang, 1998) The total Chinese population of P. b. chinensis was estimated at 1-1.5 million in the wild and the population of P. b. bengalensis at 500,000-600,000 (Anon., 1985 in CSIS, 1998). This estimate was calculated by Professor Tan Bengije and Dr. Du Houji (IUCN Cat Specialist Group) in 1985 using annual hide records (Jackson & Nowell, 1996). Wang (1998) stated that no more information on populations size has been available since 1985. The southern regions of China (Yunnan, Guizhou, Guangxi and Sichuan) were reported to be home to the highest P. bengalensis populations (Wang, 1998). More recently, a survey using camera traps and informant questionnaires to find signs of wild South China tigers Panthera tigris amoyensis in eight reserves spanning five provinces of South-central China found that Leopard Cats were present in only one reserve out of these eight reserves: the Hupingshan National Reserve, Hunan Province (Tilson et al., 2004). The main threat to its survival in the wild was reported to be hunting for skins; Wang (1998) suggested that since it was not listed in the China State Key Protected Wildlife List, hunting and trading were not regulated. The Leopard Cat is, however, protected at a regional level, e.g. in Shaanxi Province where it is listed as the local key protected wildlife (Zeng et al., 2007).

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The Leopard Cat is also hunted for food and is a speciality in Guangdong, Guangxi (as well as Hong Kong, and Macao) (Wang, 1998). Thousands were estimated to have been slaughtered for this purpose. Traditional Chinese medicine recognises leopard cat dried bones to have properties in ‘relieving rheumatic pains, strengthening bones and muscles, nourishing and calming’ (replacing the trade in tiger bones) and therefore may involve a notable level of harvest for this purpose (CSIS, 1998; TRAFFIC, 2000). In 1997, a survey in Yunnan Province on the border with Viet Nam found skins of Leopard Cat being sold in cities in the region (Li & Wang, 1999). Forest clearance for plantations was also reported to be a threat, resulting in a loss of natural habitats for felids (Wang, 1998). Deforestation has occurred within and around the Niubeiliang National Nature Reserve, Shaanxi Province; however, habitat restoration work is apparently underway in order to enlarge the distribution range of animals (Zeng et al., 2007). Leopard Cats may also be affected by rodent poison, which can kill them when ingested (Wang, 1998). In 1985, Chinese populations of subspecies P. b. bengalensis were transferred from Appendix I to II of CITES, and subsequently exports increased and caused a reduction of the population (Wang, 1998). An estimated 800,000 P. b. bengalensis were killed in Yunnan and Guizhou between 1985-1988 alone, and up to 1.87 million Leopard Cats were killed in total between 1985-1989, including many of the more valuable P. b. bengalensis skins; trade in juvenile skins predominated (Yu Jinping, 1991 in WCMC & IUCN/SSC Trade Specialist Group, 1992). Trade in Leopard Cat skins in 1989-1990, was only one half or one third of 1960s levels. The CITES Standing Committee receommended all Parties to suspend imports Prionailurus bengalensis from China in 1993 (CITES Notification No. 737) until recommendations for the conservation of this species had been implemented, as identified through the Significant Trade Review process. The Standing Committee withdrew its recommendation to suspend imports in 1995 (CITES Notification No. 869). The CITES Secretariat received information stating there has been no legal taking of specimens of Prionailurus bengalensis from the wild in China since 1989, and that such taking is still prohibited (Notification No. 869). The Management Authority of China did not permit the export of specimens of this species from April 1993, although a stock of 803,052 skins, plates and garments derived from skins of this species remained. Each item was marked by the Management Authority of China with a tag attached before export with a unique serial number and a running total of the number of skins whose export has been authorized (Notification No. 869). The International Society on Endangered Cats (2001) reported that the Chinese Government had proposed that their legal annual quota of Leopard Cat skins be increased by 500%, particularly in view of the economic value that can be derived from this national natural resource. However, no field studies have been carried out on this species in China, and its population density is unknown (ISEC, 2001). The population of Leopard Cat (P.b.bengalensis) was described as ‘satisfactory’ in 1985 by Professor Tan Bengije and Dr. Du Houji (cited in Anon, 1985). However, another member of the Specialist Cat Group, Paul Leyhausen, stated there was insufficient information on the distribution of P.b.bengalensis and P.b.chinensis to warrant a change in CITES Appendix listing (cited in Anon, 1985). The Leopard Cat was listed as Vulnerable in the Chinese Red Data Book (Wang, 1998). Wang (1998) recommended that the leopard cat should be listed in national legislation (Second Category of State Key Protected Wildlife List) to Ensure its protection and sustainable use. They also suggested that more field surveys and monitoring should be carried out on population status and distribution and that a reasonable harvesting quota should be set up, based on biological information.

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CHINA, HONG KONG: P. b. chinensis has been reported here; however, it was reported as rare by Marshall (1967). The subspecies was considered to be ‘probably absent from Hong Kong Island’ with ‘a few probably still remaining in the New Territories and close to Fanling’ (Marshall, 1967). Many medium and large mammals, including the Leopard Cat, were thought to be close to local extinction during the 1960-1970s (Lofts, 1976). AFCD (2006) reported that reforestation in Hong Kong's protected areas had meant that most populations were successfully restored and were no longer at risk of local extinction. The Leopard Cat is widely distributed in forested areas throughout Hong Kong, except Lantau Island (AFCD, 2006). The Leopard Cat was considered a culinary delicacy and hunted at least in the past (Wang, 1998). The Leopard Cat is currently protected under Hong Kong law ‘Wild Animals Protection Ordinance Chapter 170 and the Animals and Plants (Protection of Endangered Species) Ordinance Chapter 586 (AFCD, 2006). Under the Wild Animals Protection Ordinance Chapter 170, hunting of any protected wild animal is prohibited, except in accordance with a special permit (AFCD, 2006). It ws considered ‘uncommon’ by the Agriculture, Forestry and Conservation Department (AFCD, 2006). CHINA, MACAO: The Leopard Cat is hunted for food in Macao and is considered a speciality in Macao. As in other parts of China, many Leopard Cats are harvested for this purpose (Wang, 1998). CHINA, TAIWAN: P. bengalensis has been recorded in Taiwan (Corbet & Hill, 1992) and reported as P. b. hainana by Wang (1998). P. bengalensis was considered as one of the commonest wild cat species by Santiapillai & Supraham (1985) but it was later considered ‘uncommon on the island of Taiwan’ (Nowell & Jackson, 1996) and had become rare due to habitat loss (Nowell, 1991 in Nowell & Jackson, 1996). Nowell (1991, cited in Nowell and Jackson, 1996) reported that in Taiwan, Leopard Cats were seldom caught in the traps set by aboriginal hunters. DEMOCRATIC PEOPLE'S REPUBLIC OF KOREA: P. b. euptilura was reported to be present and has been recorded north to the Amur Basin (Pocock, 1939). INDIA: Present in India (Gurung & Singh, 1996) where it is well documented. Three races were reported to occur: P. b. horsfieldii, P. b. trevelyan, P. b. bengalensis. The Leopard Cat is distributed in the forest regions of India from Kashmir and the Himalayas to Cape Comorin (Prater, 1965). A survey of relatively unexplored high altitudes (2,500–5,000 m) of western Arunachal Pradesh, in north-east India, situated in the transition zone between the Himalayan and Indo-Burmese regions, reported the presence of Leopard Cats (Mishra et al., 2006). It has been described as well distributed in forests, plains and hills and still fairly common, being the second most common small felid in north-eastern India after the Jungle Cat (Felis chaus) (Choudhury, 2003). It is the most common small felid within the forest area (Choudhury, 2003). The presence of Leopard Cats has been confirmed in at least 15 protected areas in Assam, 11 protected areas in Arunachal Pradesh, 4 protected areas in Manipur, all four protected areas of Meghalaya, all 6 protected areas of Mizoram, 11 protected areas of Nagaland, 4 protected areas of Sikkim and 3 protected areas in Tripura (Choudhury, 2003). Its presence has been confirmed in the Amchang Wildlife Sanctuary near Guwahati, the capital of Assam (Choudhury, 2005). A local report confirmed the Leopard Cat as common in the Dibru-Saikhowa Sanctuary, in eastern Assam (340 km²), where tree felling was the greatest problem, and poaching was only an issue when floods forced animals into higher ground close to villages and tea estates (Choudhury, 1998). Leopard Cats were also found to be present in the Western Ghats, Kalakad-Mundanthurai Tiger Reserve (Mudappa, 2002). A comparison of two rainforest areas, Brahmagiri-Makut and Sirsi-Honnavara of the Western Ghats, Southern India found that Leopards Cats were recorded only in Brahmagiri-Makut at a relatively low rate frequency of 0.037/hour (Kumara & Singh, 2004).

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The Eastern Ghats (11°30’ - 22°N and 76°50’- 86°30’E) which are spread over the states of Orissa, Andhra Pradesh, and Tamil Nadu, support small populations of Leopard Cat, especially in specific niches of the remaining pockets of jungle (Shekhar, 2005; Srinivasulu, 2005). The study by Srinivasulu, 2005 found that local hunters considered Leopard Cats to be a chance catch compared to other wild cats in the area. The main threat in this region facing the Leopard Cat was habitat destruction and fragmentation with information on exact distribution lacking. In Andhra Pradesh the species has been classified as Least risk/near threatened; it has a very patchy distribution with populations rapidly dwindling due to habitat loss and killings (Srinivasulu, 2005). An assessment of people’s consumption of wild cat meat found locals in the Eastern Ghats perceived the Leopard Cat to be a ferocious animal and a culinary delicacy if captured (Kolipakar, 2004). The same study used camera traps; only two Leopard Cats were sighted and local people said that the species was very rare but was sometimes found in thick forest. It was concluded that not enough was known on the impact of consumption of wild cats in the region (Kolipakar, 2004). Leopard Cats were killed for their skins in western Arunachal Pradesh, and these were reported to be traded locally with hunting being motivated by poultry depredations, sport, and skins (Mishra et al., 2006). The Leopard Cat is listed at the highest protection and conservation level on Schedule I of the Wildlife (Protection) Act 1972, which prohibits all killing and capture (Choudhury, 2003). There are, however, problems with enforcement in remote areas (Choudhury, 2003). INDONESIA: The Leopard Cat has been recorded in Kalimantan (Medway, 1977; Payne et al., 1985), Sumatra, Java, Bali (Medway, 1978; Payne et al., 1985; Santiapillai & Supraham, 1985) and perhaps Lombok (Corbet & Hill, 1992). Several subspecies are considered to occur in Indonesia; P. b. borneoensis in Kalimantan; P. b. javanensis in Java and Bali and P. b. sumatranus in Sumatra (Wilson & Reeder, 2005). The presence of Leopard Cats was confirmed in the Way Kambas National Park in the Lampung Province in south-eastern Sumatra using both ecological surveys and thorough interviews with local people (Nyhus et al., 2003). Despite being one of the most common small cat species, the Leopard Cat was seldom photographed in a three year study using camera traps in the Kerinci Seblat National Park, Sumatra (1°07’S/100°59’E, 3°43’S/102°89’E) which covers 13,300km², the largest protected area in Asia (Holden, 2001). A study by Santiapillai & Supraham (1985) stated that it was not likely to be poached in Sumatra where its flesh is not sought after. The main threats in Sumatra were considered to be habitat alteration and intensive forest clearance (Santiapillai & Supraham, 1985). In certain parts of Sumatra, Leopards Cat cubs were found for sale as pets, including in the Province of Bengkulu for $10 (Santiapillai & Supraham, 1985; Holden, 2001). These were usually cubs found orphaned after their parents have been killed for raiding livestock (Holden, 2001). Leopard Cats are generally not killed for pelts or for meat in Indonesia, particularly as it is prohibited for Muslims to eat wild animals such as primates or wild cats (Holden, 2001). However, this may not be the case on all Indonesian islands where other cultures/religions are prominent, such as in Borneo where Dayak people eat forest animals (pers. obs.). It is widely distributed throughout Sumatra except for the island of Nias (Santiapillai & Supraham, 1985). A nine month study in Sumatra (Scott et al., 2004) confirmed other studies undertaken in Malaysian Borneo (Sunquist & Sunquist, 2002; Rajaratnam et al., 2007), which indicated that the Leopard Cat is capable of surviving and adapting relatively well to fragmented, logged and oil palm plantations. Leopard Cat prints were found more frequently in forest and scrub habitats (1.6/km) than in oil palm plantations (0.6/km) but encounter rate was greater in oil palm (0.24/km) than in scrub (0.12/km of forest ecotone (0.04/km) from a total of 76 Leopard Cat sightings (Scott et al., 2004). 135 photographs of Leopard Cats were captured from over 4000 camera trap nights (Scott et al., 2004).

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The species has been a protected species since 1973 (Santiapillai & Supraham, 1985) under Regulation No.66/kpts/Um/2/1973 when it was added to the colonial Dutch regulation from 1931, as a protected and ‘endangered’ species native to Indonesia. Other legislation outlines the rules of all protected species included in this regulation. The Protected Wildlife Inventory (No. 301/kpts/1991) states that all parts of listed animals and plants, whether dead or live, require a permit to allow them to be kept, which is delivered by the Natural Resources Conservation Bureau. There are guidelines for the wild-take of protected species (No. 104/kpts-II/2000), which set out rules for hunting permits. A clause in the National Regulation for the preservation of flora and fauna (No.7, 1999) allows ‘dangerous’ animals to be killed by a designated government official. JAPAN: P. b. euptilura occurs only on the Tsushima Islands, which have an area of approximately 708 km² . It is considered endangered because of its reduced population size, down to about 70 in the 1990s (Cat News , 2000; Izawa & Doi 1991; Izawa et al., 1991), from perhaps 200-300 individuals in the 1960s-1970s (Nowell & Jackson, 1996). KOREA, REPUBLIC OF: P. b. euptilura was common before World War II and the Korean War but then considered rare (Won, 1976). LAOS PEOPLE'S DEMOCRATIC REPUBLIC: Osgood (1932) originally recorded the species from Ban Muangyo, Phongsali, Ban Thateng, Ban Phonsavan, while Deuve (1972) considered the Leopard Cat to be present in all provinces of Laos, where it occurs as P. b. bengalensis (Osgood, 1932; Ellerman & Morrison-Scott, 1966). The political climate throughout the 1970-1980s prevented foreign scientists undertaking any survey of the flora and fauna until about 1988 (Duckworth, 1997). However, in the late 1990s, when Duckworth (1997) conducted a small carnivore survey, he noted the Leopard Cat to be ‘the only small cat species observed frequently in Laos’. The same study also found that the live Leopard Cats and derivatives (pelts) were generally not found in market places (other small carnivores appeared regularly) but live captives were seen in menageries in Ban Lak and Vientiane Cultural Park. A few skulls were seen and pelts were seen hanging on walls, usually following poultry farm raids by these carnivores. Locals in Sangthong did not like to eat Leopard Cat meat but it was freely eaten in other regions of Laos such as Nam Kading. The Leopard Cat was the most frequently sighted (12 direct sightings) in Duckworth’s (1997) study from 1992-1996, with several sightings in degraded forest areas and on logging roads or in scrub. Duckworth (1997) recorded >25 sightings of Leopard Cats (including a range of survey methods observations in villages, using camera traps in addition to direct sightings) in Lao PDR and he indicated that it was by far the most commonly found small or medium sized cat in comparison to other small felids (Duckworth et al., 2005). The two major threats to the Leopard Cat and other small cats were cited by Duckworth (1997) as habitat loss and direct harvesting usually related to opportunistic hunting by humans. Article 5 of the Decree on the management and protection of wild animals, fisheries and on hunting and fishing (Council of Ministers, 1992) declared that all Lao citizens have the right to capture or kill restricted animal species if human lives are in danger, following which the captured or killed animals become the property of the state. In cases where an animal species is causing damage to livestock or plantations, it must be reported to the local forestry and environment management authorities and provided with an adequate solution. MALAYSIA: Occurs as P. b. borneoensis in Sabah and Sarawak (Medway, 1977) and as P. b. sumatranus in most habitats in Peninsular Malaysia (Medway, 1978). It is found in most areas in Borneo, including in some towns (Medway, 1977; Payne et al., 1985). These records include parts of the mainland, lowlands and from high ground (up to 2,000 - 3,800 feet) on G. Dulit, the Kelabit uplands and on Kinabulu (Medway, 1977). It was considered the only small wild cat that has been considered common in Malaysia (Tweedie, 1978). The Leopard Cat was recorded as one of the most frequently photographed species in a study using camera traps to assess wild felid populations in the Jeranau Forest Reserve, Ulu

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Terengganu (140, 000ha; 4º55.5’N, 103º5.7’E) in Peninsular Malaysia (Azlan & Sharma, 2006). Although it was relatively common in this area, this Forest Reserve was considered to be a wild felid hotspot and may not represent population levels elsewhere. Another study in Peninsular Malaysia’s most important conservation area, Taman Negara National Park, (4º10’–4º56’N, 102º00’–103º00’E) photographed Leopard Cats using camera traps on 59 occasions (Kawanishi & Sunquist, 2004). The major threat was reported to be logging, even in the Forest Reserve, which may not stay protected in the future (Alzan & Sharma, 2006). These authors did, however, find that the Leopard Cat and other wild felids may be able to persist in secondary forest. The presence of Leopard Cats in the Tabin Wildlife Reserve in eastern Sabah was recorded during a study carried out in 1995, as well as in adjacent logged dipterocarp forest, oil palm plantations and secondary forests (Rajaratnam et al., 2007). A notable result from this study, and similar studies in Sumatra (Scott et al., 2004), was that Leopard Cats seemed to exhibit a significantly high use oil palm areas where their preferred prey (murids) thrive (Rajaratnam et al., 2007). Large areas of privately owned forests are being cleared rapidly for oil palm plantations throughout Borneo (both in Sabah and Kalimantan), potentially reducing breeding habitat for the Leopard Cat (Rajaratnam et al., 2007). It is protected by law under Schedule One of the Protection of Wildlife Act 1972, which prohibits hunting and trade (Azlan & Sharma, 2006). MYANMAR: Occurrence of P. b. bengalensis reported(Pocock, 1939; Ellerman & Morrison-Scott, 1966). The Leopard Cat is present in the Hlawga Wildlife Park, 35 km north of Yangon (17º02’35”N, 96º06’49”E), one of Myanmar’s smallest protected areas (6.24 km²) (Su, 2005). A wild cat survey undertaken in 2000, interviewing villagers, fisherman and hunters on the islands of the Myeik archipelago and coastal Tanintharyi in southern Myanmar, confirmed the presence of relatively high populations of Leopard Cats in the latter region (Jorio, 2000). Strip transect and camera trap surveys in Hkakaborazi National Park, north Myanmar recorded a

relative abundance of 3.95 (capture rate/100 trap nights) (Rao et al., 2005). A previous study, carried out between 1996-1997 in North Myanmar (24-28ºN and 97-99ºE), a narrow strip along the western escarpment of the Yunnan Province in China, reported a large number of mammal species in the area but did not report the presence of the Leopard Cat (Rabinowitz & Khaing, 1998). Hunting for subsistence and trade, together with habitat loss due to logging and shifting cultivation, adversely affect biodiversity in north Myanmar (Lynam, 2003 in Su, 2005). Staff at Hlawga Wildlife Park indicated that the Leopard Cat and other felids had previously been more common (Su, 2005). However, traps and snares were commonly set by villagers in the core and buffer areas of the Park for meat and traditional medicine. Although organised poaching and illegal trade did not seem to be in place in the coastal region of Tanintharyi, local people regularly hunt all wild felids in the region for meat and medical purposes (Jorio, 2000). Rao et al. (2005) also found P. bengalensis in meal records of local peoplein north Myanmar. Traders in eastern Myanmar, close to the Thai border, are well known for selling wild cat products (Martin, 2001). Retail buyers from Singapore, Thailand, Japan and Taiwan go there to smuggle such products over the border. Although trade is prohibited in both Myanmar and Thailand, neither government takes any enforcement action. More than 100 skins were found for sale in Tachilek, Myanmar in February 1998 at $12 each (Martin, 2000). Another threat to the Leopard Cat was reported to be deforestation and conversion into oil palm plantations (Jorio, 2000). Although biodiversity conservation is part of Myanmar’s current government policies and over 96 mammal species are listed, including the Leopard Cat which has ‘protected’ status, Su (2005) recorded indicators of intense human pressure such as timber

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removal, forest fires, illegal hunting, snaring and trapping within core and buffer areas of the Hlawga Wildlife Park. Unlike many of its neighbours, Myanmar still contains large areas of relatively intact forest, particularly north of 24ºN, but these resources are under tremendous pressure for exploitation (Global Witness, 2003 in Su, 2005). Conservation legislation in Myanmar consists of the “the Protection of Wildlife and Protected Areas Law; State Law and Order and Restoration Council, 1994” which states that the Director General may, with approval of the Minister, declare wildlife to be completely protected wild life; normally protected wild life or seasonally protected wild life, according to the danger of extinction. “Wildlife conservation legislation enacted in Myanmar in 1994 ensures protection of wild flora and fauna, their habitats and representative ecosystems. Completely Protected species may not be hunted except for scientific purposes under a special licence; (2) Protected species may be hunted but only with special permission; (3) Seasonally Protected species are subject to traditional subsistence hunting by rural communities only during the open (i.e. non-breeding) season. Hunting is illegal within protected areas but legally allowed elsewhere including, with a licence, buffer zones of protected areas. In reality, however, the Forest Department no longer issues licences, and hence hunting within buffer zones is considered an illegal activity” (taken from Rao et al., 2005). Rao et al. (2005) heavily criticized current legislation and highlight the need, from their results (which indicated that offtake of species was independent of protection status), for legislative reform and proper enforcement of national legislation. They also insisted on the fact that Myanmar has been a CITES Party since 1997 and should therefore have CITES-enabling legislation in place. This point is also emphasized by (Pant, 1998 in Rao et al., 2005), who stated that legislation relevant to wildlife conservation in Myanmar was ’inadequate, often ambiguous, and rarely properly enforced’. NEPAL: Occurrence reported in Nepal (Gurung & Singh, 1996) where it occurs as P. b. horsfieldii and infrequently as P. b. trevelyani (WCMC & IUCN/SSC Trade Specialist Group, 1992). Surveys in the late 1980s and 1990s by conservation organizations in Kathmandu, Nepal, uncovered a major illegal trade in protected animals including 60 Leopard Cat coats on open display (Heinen & Leisure 1993). A follow−up visit to Kathmandu in 1993 found no change; protected species continued to be sold openly (Menon, 1994). It is protected under the Protected Animal Species and National Parks and Wildlife Conservation Act, 1973. PAKISTAN: Occurs in North Pakistan (Gurung & Singh, 1996) in the region of Baluchistan as P. b. trevelyani (Ellerman & Morrison-Scott, 196; Nawaz, 1983) and P. b. horsfieldii has been reported from Islamabad (Chakraborty, 1983). Normally associated with moist temperate forests of the Himalayas but can also be found in dry coniferous forests of the Himalayas further north, as well as dry sub-tropical forests on the foothills. Although it has become extremely rare in areas where it was previously more abundant (Nawaz, 1983; Roberts, 1997), due to persecution for its fur, it can still be found in the Murree Hills and forests of the Neelum Valley in Azad Kashmir, as well as the Kaghan Valley in Shoghran (Roberts, 1997). It has also been reported in the Margalla Hills at Chak Jabri, although its occurrence there is rare. It is also known from Swat, Kohistan, lower Gilgit, northern Dir forests and lower Chitral, although it was thought to be quite rare, particularly in the last two regions (Roberts, 1997; Wildlife of Pakistan, 2004). It is most probably absent from more southern regions and, although skins have been reported from these regions (including a specimen at the British Natural History Museum), these were thought to probably be unusually marked Felis silvestris (Roberts, 1997). Persecution of the Leopard Cat and other carnivores (in response to livestock/ crop losses) is common in Pakistan (Nawaz, 1983; Wildlife of Pakistan, 2004).

Although the leopard cat is not listed nationally as a species of concern (Wildlife of Pakistan, 2004) it is present in the following National Parks: Khunjerab 226,913 ha; Central Karakorum 1,390,100 ha; Deosai Plains 358,400 ha; and Chitral Gol 7,750 ha.

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Roberts (1997) stated that the Leopard Cat was in danger of becoming extinct in the Pakistan region if no restrictions on the pelt trade were put it place. They fetch a high price and hunters usually report skins as being from the regions of Dir and Swat (Roberts, 1997). PHILIPPINES: It is found in Palawan Island (Payne et al., 1985): one was spotlighted in 2004 (Esselstyn et al., 2004). It was also reported to be present in the neighbouring island of Busuanga, as well as in Cebu, Negros and Panay (Corbet & Hill, 1992; Heaney et al., 1998; Heaney et al., 2002). The population of Palawan Island faunal region was described as a separate subspecies, P. b. heaneyi and Leopard Cats in Negros, Cebu, and Panay were described as P. b. rabori (Groves, 1997). Island populations were seriously threatened in the Philippines and both subspecies were listed as Vulnerable in the 1997 Philippines Red Data Book (Nowell & Jackson 1996). Uncommon but widespread; heavily hunted and declining in the Philippines (Heaney et al., 2002). RUSSIAN FEDERATION: Found in the Amur Valley, Amur-Ussuri (far east Siberia) on the border with China (Ellerman & Morrison-Scott, 1966). It was proposed by Geptner (1971) that the Russian population and part of the Chinese one are a separate species, P. euptilura. Many Russian authorities refer to the Leopard Cat as P. euptilura, which is usually slightly larger than the Asian Leopard Cat (Roberts, 1997). Currently, it is still considered a subspecies, P. b. euptilura (Wilson & Reeder, 2005). There is concern about the status of the species in the Russian Far East (Nowell & Jackson 1996). SINGAPORE: The subspecies, P. b. sumatranus has been recorded in Singapore (Harrison, 1974; Medway, 1978). The last confirmed sighting was from the Mandai area in 1968 and it may be on the brink of extinction (Ng & Wee, 1994). The main threat has been the ‘shrinkage’ of forests and habitat fragmentation. There was a plan to build corridors between the Bukit Timah Nature reserve and the central catchment area in 1994 (Ng & Wee, 1994). THAILAND: The Leopard Cat has been considered to be one of the commonest wild cats in Thailand and the least threatened Thai felid (Lekagul & McNeely, 1977; Humphrey & Bain, 1990, Grassman et al., 2005). It occurs throughout Thailand including on many offshore islands. Present as P. b. bengalensis (Osgood, 1932). Leopard Cats occur in many protected areas (WARPA, 1980) A study from the Huai Kha Khaeng Wildlife Sanctuary in west central Thailand found that home range varied between 1.5-7.5 km² with core areas of 2 km² (Rabinowitz, 1990). Its habitat preference was for evergreen or mixed deciduous forests rather than dry dipterocarp deciduous forests (Rabinowitz, 1990). Its presence is also confirmed in the Phu Khieo Wildlife Sanctuary in north-central Thailand (16º5’-16º35‘N, 101º20’- 101º55’E) which is one of the three Thai protected areas that does not contain a permanent human settlement, although illegal hunting, aloewood extraction and tourism may have some level of impact (Grassman et al., 2005). Mean home ranges were found to be significantly higher than in any other study to date, at 12.7 km², compared to mean home ranges from 1.5-8 km² in other areas of Thailand (Grassman et al., 2005). Thai Law - Wild Animals Reservation and Protection Act of 1960 - prohibits hunting and regulates trade in the species (Rabinowitz, 1990). VIET NAM: The Leopard Cat occurs as P. b. bengalensis (Ellerman & Morrison-Scott, 1966). No comprehensive overview of survey results has been published, which means that the total number of records of felids during the 1990s is impossible to calculate. Pham Nhat (verbally 1997) reported that it was easy to find stuffed examples of Leopard Cat, and other species, in Vietnamese taxidermists (Duckworth et al., 2005). In an assessment of mammalian populations in Cat Tien National Park and its surrounding areas in southern Viet Nam, only the Leopard Cat’s presence was confirmed out of the five possible species of cats (Polet & Ling, 2004).

25

The Leopard Cat was included in Group II of Viet Nam’s list of rare and precious flora and fauna and regulations for their management and protection in 1992, which was updated in 2002). This classification means that it is part of a list of national species that have “a high economic value and are over-exploited, leading to a state of depletion and danger of extinction” (Decree No.18/HDBT and DECREE No. 48/2002/ND-CP).

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REPTILIA: ALLIGATORIDAE

SPECIES: Alligator mississippiensis

SYNONYMS: Alligator helois, Crocodilus cuvieri, Crocodilus lucius,

Crocodilus mississippiensis COMMON NAMES: Mississippi Alligator (Dutch), American Alligator

(English), Alligator américain (French), Alligator du Mississippi (French), Aligator del Mississippi (Spanish), mississippi alligator (Swedish)

RANGE STATES: United States of America RANGE STATES UNDER REVIEW: United States of America IUCN RED LIST: Lower risk/least concern PREVIOUS EC OPINIONS: A positive opinion was formed for the USA on

18/07/2001, and confirmed on the 29/10/2001

TRADE PATTERNS: The American Alligator, Alligator mississippiensis, was selected for this review on the basis of high volumes of trade in 2004; it also showed a sharp increase in trade of wild sourced skins into the European Community in 2004 (UNEP-WCMC, 2006). The American Alligator has been heavily exploited since the 1800s. The high-quality leather produced from the belly skin of the alligator resulted in considerable hunting pressure in the first half of the 20th century, until the species was designated as endangered in the 1960s; trade resumed in the early 1970s after a rapid recovery as a result of a good response to management (USFWS, 1987; Ross, 1998; Britton, 2007). The volume of reported trade in alligator skins has increased steadily, although with some fluctuations, during the last 30 years: from around 20,000 skins reported in trade in the early 1980s, to over 150,000 in the early 1990s, and to around 350,000 in the early 2000s (UNEP-WCMC, 1994-2006).

Alligator skin prices have been fluctuating during the last decades, and wild-sourced skins have generally been more highly valued than ranched ones. The value index of wild-sourced alligator skins rose from US$ 2-3/cm in the early 1980s to over US$ 9/cm in the late 1980s, then falling to US$ 4/cm in the early 1990s and rising again to US$ 7/cm in 1995, only to fall once more to around US$ 3/cm in the late 1990s; skins from ranched alligators have followed a similar pattern, although prices reached peaks of only over US$ 6/cm in the late 1980s and of US$ 4.80/cm in 1995, while prices during the low-price periods were about the same as those for wild-sourced skins (Woodward & David, 1993; MacGregor, 2002).

In the early 1980s, almost all alligator skins supplied to the industry were wild-harvested. During the following years, a growing proportion of skins reported in trade were from ranched and captive-bred alligators, and in the last decade (since the mid-90’s) only around one-quarter of skins traded were reported as wild-sourced (MacGregor, 2002; Caldwell, 2006). However, there have been some problems with the source codes used for alligators by the United States, and ranched specimens have apparently been reported as either captive-bred

31

or wild-taken; the sharp increase in reported trade in wild-sourced skins in 2004 was likely to have been as a result of ranched specimens being reported as wild-taken.

The EU is a major importer of alligator skins, with France, Germany and Italy being the main importers. They account for 95% of EC-reported imports and over 50% of global trade (UNEP-WCMC, 2006); together with Singapore they import 95% of the world’s production (Caldwell, 2006). Nearly all imported skins are from the US, the only range state. A summary of CITES trade data is provided in Tables 8-10. No quotas have been published for Alligator mississippiensis.

32

Table 8. Main direct exports of Alligator mississippiensis to EU-27, 2001-2005.

Exporter Term Source Reported by 2001 2002 2003 2004 2005 Total

Israel live C Importer 63 9 105 151 328

Exporter 63 9 94 94 260

skins C Importer 239 239

Exporter 6 6

Mexico skins C Importer

Exporter 5 5

South Africa live C Importer 5 5

Exporter 5 5

Ukraine live C Importer 9 9

Exporter

United States bodies U Importer 6 6

Exporter

W Importer 10 10

Exporter

garments C Importer 2 1 3 2 8

Exporter 4 17 2 1 24

R Importer

Exporter 3 3

W Importer 5 1 7 13

Exporter 8 11 2 2 23

large leather products C

Importer 42 16 18 6 16 98

Exporter 19 3 50 18 20 110

F Importer 1 1

Exporter 1 6 7

R Importer 4 1 2 12 19

Exporter 5 1 8 17 1 32

W Importer 8 21 32 24 47 132

Exporter 58 175 12 30 30 305

- Importer 79 2 81

Exporter

live C Importer

Exporter 30 30

F Importer 1 1

Exporter 31 31

W Importer 16 16

Exporter 16 16

meat (kg) C Importer 1088.4 1088.4

Exporter 1362 817 2179

F Importer 1089 1089

Exporter 1632 2177 3809

shoes C Importer 7 95 105

Exporter 67 31 445 630 141 1314

F Importer

Exporter 4 1 5

R Importer 3 6 9

Exporter 10 12 20 6 12 60

W Importer 2 15 17

Exporter 6 33 37 76 70 222

33

Exporter Term Source Reported by 2001 2002 2003 2004 2005 Total

skin pieces (kg) C

Importer

Exporter 68 68

R Importer 238 238

Exporter 102 276 378

skin pieces (unspecified units) C

Importer 133 3460 2903 6496

Exporter 1 10981 68 38 18 11106

F Importer

Exporter 7010 12516 19526

R Importer 102 50 152

Exporter 750 1243 10 2003

W Importer 604 604

Exporter 2326 2367 95 277 276 5341

skins C Importer 181765 163843 168038 89640 47 603333

Exporter 151553 91605 146454 149176 538788

F Importer 3072 982 3083 1108 8245

Exporter 37237 23719 19117 635 80708

R Importer 40000 14933 45530 20573 716 121752

Exporter 40614 21536 48756 18974 811 130691

W Importer 40940 30881 34185 70167 329027 505200

Exporter 38472 29651 32654 101285 254640 456702

skulls C Importer 4 2 6

Exporter

U Importer 3 6 2 11

Exporter

W Importer 8 2 1 11

Exporter 7 2 1 4 23 37

small leather products C

Importer 658 3111 797 3048 1407 9019

Exporter 144 358 172 3315 5573 9562

F Importer 4 6 11 2 1 24

Exporter 1535 255 2 1792

I Importer 7 28 84 119

Exporter

R Importer 13 80 33 11 19 156

Exporter 13 1 14

W Importer 1077 52 126 462 1826 3541

Exporter 560 197 100 511 371 1739

tails R Importer

Exporter 929 929

trophies I Importer 1 4 1 6

Exporter

W Importer 5 20 25

Exporter 5 10 15

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Table 9. Main indirect imports of Alligator mississippiensis to EU-27, 2001-2005.

Origin Term Source Reported by 2001 2002 2003 2004 2005 Total

Colombia small leather products C Importer 1 358 133 150 10 652

Exporter 96 234 330

- Importer 146 146

Exporter

United States bodies W Importer

Exporter 105 105

garments, shoes (pairs) and large leather products C

Importer 762 223 181 59 70 1295

Exporter 19 8 3 52 0 82

R Importer 17 0 0 25 32 74

Exporter 2 1 0 6 0 9

W Importer 529 207 152 75 123 1086

Exporter 4 3 3 6 60 76

live C Importer 10 15 9 34

Exporter 12 15 27

F Importer 1 6 7

Exporter 6 6

W Importer 15 4 19

Exporter 2 15 17

skins (bellyskins) C Importer

Exporter 1656 310 200 2166

W Importer

Exporter 129 40 1020 1189

skins (sides) C Importer 4 344 200 5200 5816 11564

Exporter 2341 620 200 5044 8205

R Importer 3 3

Exporter 3 3

W Importer 80 80 2040 2200

Exporter 258 80 2250 2588

skins (unspecified units) C

Importer 26907 28164 31565 19425 23908 129969

Exporter 24076 30003 30410 35933 12333 132755

F Importer 3 119 122

Exporter 50 128 178

R Importer 561 10 474 196 247 1488

Exporter 92 26 476 233 28 855

U Importer

Exporter 125 125

W Importer 6570 2225 3368 4836 21384 38383

Exporter 3263 2244 2973 4092 22952 35524

- Importer

Exporter 3744 3744

small leather products and skin pieces C

Importer 231042 235614 310317 362352 423862.1 1563187

Exporter 285558 302660 346807 460728 72596 1468349

F Importer 5136 6453 2435 647 2361 17032

Exporter 7794 6658 2513 1023 1799 19787

I Importer 60 687 747

Exporter

35


R Importer 16438 14892 15906 14347 38720 100303

Exporter 16447 20278 19592 31326 12786 100429

U Importer 70 26 45 7 148

Exporter 2 40 92 134

W Importer 3319 5740 9033 7495 54746 80333

Exporter 5040 6109 12489 19145 24824 67607

- Importer 28 5537 3738 4899 4556 18758

Exporter 5 5

tails C Importer 40 40

Exporter

W Importer 145 20 165

Exporter 145 145

trophies C Importer 6 6

Exporter

I Importer 3 3

Exporter

Table 10. Main direct exports of Alligator mississippiensis to countries other than the EU-25, 2001-2005.

Exporter Term Source Reported by Total

Colombia skins C Importer 1000

Exporter

Italy skin pieces C Importer 20

Exporter

skins C Importer 70

Exporter

Mexico small leather products (pairs) C

Importer

Exporter 55

United States bodies (kg) F Importer

Exporter 465

bodies (unspecified units) I

Importer 5

Exporter

W Importer 40

Exporter

garments, large leather products and shoes C

Importer 27

Exporter 13906

R Importer 1

Exporter 190

W Importer 14

Exporter 588

live C Importer 49

Exporter 12

F Importer 68

Exporter 53

meat (kg) C Importer 22532.88

Exporter 25670

F Importer 1477.04

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Exporter 3416

R Importer 54.43

Exporter 1190

W Importer 115800.64

Exporter 148160

skin pieces and small leather products C

Importer 59023

Exporter 89161

F Importer 1577

Exporter 2776

R Importer 670

Exporter 2340

W Importer 22727

Exporter 31087

skins C Importer 261080

Exporter 260096

F Importer 2254

Exporter 13478

R Importer 8743

Exporter 8537

W Importer 156055

Exporter 158488

- Importer 2293

Exporter

skulls C Importer

Exporter 34

W Importer 6

Exporter 2

trophies W Importer 3

Exporter 16

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CONSERVATION STATUS IN RANGE STATES: UNITED STATES: Alligator mississippiensis is known to occur in the south-eastern United States: Alabama, Arkansas, Florida, Georgia, Louisiana, Mississippi, North Carolina, Oklahoma, South Carolina and Texas (NatureServe, 2006; Britton, 2007; USFWS, 2007). A total population of around 800,000 was reported in 1982 (Groombridge, 1982). More recently, it was estimated to be over 1 million by some sources (Natureserve, 2006; Britton, 2007), although considerably higher numbers (over 3 million) can be inferred by adding population estimates of single states (e.g. Woodward & Cook, 2000; GDNR, 2007; LDWF, 2007a; TPWD, 2007a). An alligator survey in Alabama, Arkansas, Georgia, Mississippi, and South Carolina indicated the population increased by 110% between 1974 and 1982 (Chabreck, 1984, cited in USFWS, 1987). In the 1990s, alligator populations were considered to be quite healthy (Ross, 1998). NatureServe (2006) considered the alligator to be vulnerable only in North Carolina and under review in Oklahoma (marginal areas of its range), but secure or apparently secure in all other States where it occurs. The principal threats facing the American alligator were considered to be habitat destruction and environmental contamination (e.g. Ross, 1998; Jagoe, et al., 1998; Natureserve, 2006; Britton, 2007). Trade was considered to be sustainable and therefore not a main threat to the species (e.g. Ross, 1998; CSG, 2005). In the past, overexploitation and habitat destruction led to the reduction of American alligator populations, so that it was classified, in 1967, by the US Department of the Interior as an endangered species throughout its entire range (USFWS, 1977). The alligator subsequently recovered in many parts of its range in response to Federal and State protection (Groombridge, 1982), and in 1987 it was reclassified to “Threatened due to Similarity of Appearance” (as it is similar in appearance to an endangered taxon – the American crocodile Crocodylus acutus) throughout its whole range, under provisions of the Endangered Species Act of 1973 (USFWS, 1987). This listing status has been maintained to date (USFWS, 2007). Alligators are harvested primarily for their skins, although the meat is also consumed. Sustainable management programmes have been in place since the mid-1980’s, involving a combination of captive-breeding, ranching and direct cropping from the wild (Ross, 1998; Britton, 2007). There are over 150 large-scale farms in several states, most of them in Florida, Louisiana and Texas, where alligators are captive-bred and ranched (CSG, 2005; Britton, 2007;) using practises that have been subject of much research (e.g. Joanen & McNease, 1987a; Woodward, 1987; Luxmoore, 1992; MacGregor, 2006). Ranching was considered to be a highly precautionary and biologically “safe” method of harvesting (Hutton et al., 2001; MacGregor, 2006). In 2005, the Crocodile Specialist Group noted that “utilization in all States is based upon regular and intensive surveys, population estimates and long-term population trends, and the status of the species is absolutely secure throughout its range” (CSG, 2005). The increasing alligator populations and the also expanding human population in alligator habitat resulted in the establishment of nuisance alligator programmes in Louisiana, Florida, Texas, South Carolina and Georgia, where nuisance alligators can be killed or removed by permitted hunters; other states use state wildlife officials to remove problem animals (Woodward & David, 1994). Alabama: In Alabama, the alligator was considered to be “fairly common in many Coastal Plain lakes, rivers, and associated aquatic habitats, and common in Mobile-Tensaw Delta” (ADCNR, 2007a). In 2006, the first regulated alligator hunting opened in Alabama for a defined area of Mobile and Baldwin counties, with a limit of 50 alligators (ADCNR, 2007b). A total of 40

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alligators (30 male and 10 female) were harvested; wildlife biologists were reported to be analyzing harvest data to determine when and where future Alabama hunts may take place (ADCNR, 2007c). Arkansas: In Arkansas, the alligator was re-introduced in the 1970s (over 2,800 sub-adult alligators were released from 1972 to 1984), and since then numbers have increased in some areas especially in the south-west and south-east parts of the state (Barkley, 1998; AGFC, 2007). In 2004, an average of 2.3 alligators per mile was encountered along population survey routes, and in 2007 an alligator hunting season was proposed for the first time (AGFC, 2007). Florida: In Florida, the alligator was classified as a “species of special concern” (FFWCC, 2004). Night-light alligator surveys of several Florida wetlands indicated that, between 1974 and 1989, population densities increased annually by 3.5 – 7.5% for different size classes (Woodward & Moore, 1990). However, Woodward & Moore (1995) estimated that the population increased by 1.9% annually, totalling 41% during the period 1974-92. No accurate estimate of the total number of alligators present in Florida was available (Hines & Abercrombie, 1987; Woodward & Moore, 1995; Woodward, 1996); however, the population was crudely estimated at approximately 1 million (Woodward & Cook, 2000). The alligator inventory for all of Florida’s alligator farms totalled 112,000 animals in 2005 (FFWCC, 2006a). In Florida, alligators are managed on designated waterbodies, known as Alligator Harvest Management Units (Woodward, 1996; Dutton et al, 2002). Wild alligators may be harvested from public waters (“statewide harvest”), private lands, or as part of a nuisance alligator control program (Woodward, 1998; Dutton et al, 2002). For public waters, quotas are set annually. For the period 1988-2001, these annual quotas have ranged from 2500 to 4000 alligators, from 300 to 5000 hatchlings, and from 150 to 1300 nests (with between 4000 and 28000 eggs collected) (Dutton et al, 2002). On private lands, landowners may choose from several harvest options, i.e. harvesting alligators/hatchlings/eggs, the type and volume of harvest being determined based on the information available for each population. In addition, the nuisance alligator control programme is in place to permit the harvest of alligators that are determined to be a threat to the welfare of the public, or their pets or property (Woodward & Cook, 2000; Dutton et al, 2002; FFWCC, 2006a). In 1977, 708 alligators were harvested in Florida, all wild-sourced (Woodward, 1998). Since then, the number of wild alligators harvested in Florida has increased to 1,590 in 1980, 4,114 in 1985, 7,672 in 1990, 11,947 in 1995, 12,605 in 2000, and 17,064 in 2005. A high proportion (from 100% in the late 1970s to around 50% (5000-9000 animals) from the mid-1990’s to at least 2006) of these wild alligators were harvested under the nuisance alligator control program. The remainder were harvested under the public waters and the private lands harvest programs (the public waters program begun operating in 1981 and the private lands program in 1985), with around 3,000 alligators harvested annually under each program in the last decade (Dutton et al, 2002; FFWCC, 2006b). Since 1978, alligators have also been harvested from Florida farms, mostly relying on wild egg and hatchling collections to maintain alligator stock (Dutton et al, 2002; FFWCC, 2006a). Numbers of farmed alligators were small during the first decade of farming, but during the period 1990-2005, between 18,000 and 38,000 alligators were harvested annually from Florida farms (Dutton et al., 2002; FFWCC, 2006b). Between 1995 and 2005, the number of alligators harvested on Florida farms has remained fairly stable - between 20,000 and 30,000 alligators per year (FFWCC, 2006b). Harvest of wild alligators was considered to be sustainable by a number of authors (Hines & Abercrombie, 1987; Woodward & Moore, 1995; Dutton et al., 2002). Hines & Abercrombie (1987) found that the removal from the population of a substantial number of adult female alligators (93 over three years) in Orange Lake did not result in a decrease in the number of nests (before the harvest, the average annual number of nests was 71.3; the post-harvest average was 88.7).

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Rice et al. (1999) found that a harvest rate of 50% of eggs or hatchlings did not adversely affect recruitment into the sub-adult or adult size classes. Britton (2007) found that, in Florida, up to 13% of sub-adult to adult animals could be safely removed from the alligator population annually without affecting population stability. Woodward & Moore (1995) considered that illegal hunting of alligators in Florida was negligible. Like elsewhere throughout its range, habitat loss and degradation, together with environmental contamination appear to be the main threats to alligator populations in Florida (Woodward & Moore, 1995; Mazzotti et al., 2003). Georgia: The American alligator is found throughout the southern part of Georgia, up to, and occasionally beyond, the Fall Line –line along the eastern coast of the US, where the Appalachians meet the coastal plain- (GMNH, 2000). It was estimated that there were around 200,000 alligators in Georgia (GDNR, 2003; GDNR, 2006). In 2000, populations were not considered large enough to permit a regular harvest (GMNH, 2000); however, nuisance alligators have been harvested since 1989 (GDNR, 2003). More recently, the alligator population was considered to be able to sustain a limited harvest in concert with careful regular monitoring; 450 nuisance alligators are removed annually in the state (GDNR, 2007) and since 2003 regulated hunting has been permitted (GDNR, 2003). Louisiana: Louisiana’s wild population was estimated to be 26,000 in 1957 and over 300,000 in 1983 (Joanen & McNease, 1987b). More recently, the wild population in Louisiana was estimated to be just over 1.5 million animals; an additional 500,000 alligators were estimated to be kept in ranches in Louisiana (LFAAC, 2004; LDWF 2007b). The number of alligator nests in Louisiana’s coastal marshes increased steadily from less than 10,000 nests in the early 1970’s to 20,000 in 1980, 25,000 in 1990, and between 35,000 – 45,000 in 2001-2005, with 2004 being the second highest coastal nest production on record (CSG, 2005; LDWF, 2005a). From 1994-2004, between 250,000 and 400,000 alligator eggs for ranching were collected annually from the wild in Louisiana (LDWF, 2007b). In 2004, Louisiana’s alligator farmers collected 396,069 wild alligator eggs, a 9% (38,312 alligator eggs) increase compared with the 2003 season, as a result of increased alligator nest production in coastal Louisiana (LDWF, 2005a). In fact, in 2005, 507,000 eggs were collected, the highest number on record (LDWF, 2007b). Between 33,000 and 51,000 alligators have been returned to the wild annually in Louisiana during the last decade, to compensate for the egg collection for ranching (LDWF, 2005a; LDWF, 2007b). In 2004, 295,000 alligators were harvested from farms, and 33,925 alligators were harvested from the wild (LDWF, 2005a). These figures are similar to those of previous years (since 1999) for hunted alligators, but higher for farmed alligators (LDWF, 2007b). Additionally, in 2003 2,024 nuisance alligators were harvested in Louisiana (LDFW, 2005a; LDFW 2005b). Table 11. Alligators harvested in Louisiana, 1999-2005. Source: Louisiana Department of Wildlife and Fisheries. Harvested 1999 2000 2001 2002 2003 2004 2005

From farms 187,570 219,827 180,391 237,808 274,485 295,000 From wild 35,270 33,678 35,148 33,383 31,581 33,925 31,279

Nuisance animals 2,024 2,500

Hurricanes Katrina and Rita hit Louisiana in late 2005 causing physical damage to marsh vegetation and elevating the salinity levels of Louisiana marshes (LDWF, 2005c). The effect of the hurricanes, coupled with a drought in the region in 2006, negatively affected alligator nesting in some areas, particularly brackish marshes nearest the coast, with nest production in 2006 in southwest Louisiana 84.5% lower than in 2005 and in southeast Louisiana 35.3% lower (Elsey, 2006).

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Egg collection and hunting of alligators is closely regulated in Louisiana by a system of permits, tags, quotas and seasons (LDWF, 2005a; LDWF, 2005b). Louisiana’s Department of Wildlife and Fisheries must provide to the USFWS on an annual basis a “finding of no detriment” stating that Louisiana’s harvest and export of alligators are not detrimental to the survival of the species (LDWF, 2005c; LDWF, 2007b). Mississippi: In Mississippi, regulated hunting of alligators is permitted, and a total of 190 alligators can be harvested per year: 100 alligators in public waters, restricted to certain parts of the Pearl River, and 90 alligators in private lands, restricted to Hinds, Madison, and Rankin counties (MWFP, 2007). North Carolina: North Carolina represents the northernmost part of the alligator’s range. Alligators are found in the south-eastern coastal areas (ARNC, 2007) and populations are considered to be vulnerable in this state (NatureServe, 2006). Oklahoma: There were doubts in the mid twentieth century as to whether the American alligator had a breeding population in the State (Blair, 1951). Currently, it occurs only in the southeast corner of the State, particularly Choctaw and McCurtain counties (Meek, undated), where it was considered to be an occasional visitor along the Red River (USFWS, 1992). South Carolina: In South Carolina, the American alligator was listed as “a species in need of management”. Bennet & Buhlmann (2007) considered it to be common to abundant in coastal South Carolina. They noted that large, protected populations exist on some government-owned lands, such as the Savannah National Wildlife Refuge and Donnelly Wildlife Management Area. A study in Par Pond, a reservoir in South Carolina, revealed that between 1974 and 1988 the alligator population more than doubled (from an estimated 110 to 266 individuals) and that the proportion of juveniles increased considerably –by 10% per year- (Brandt, 1991). Status, distribution and trends of alligator populations are monitored annually using aerial nest surveys and night light surveys; in 1996 the general population trend of adult animals was stable or slightly increasing (Rhodes, 1996). A nuisance alligator program was established in 1988, and 238 alligators were harvested under this programme in 1996 (Rhodes, 1996). The first legal harvest of wild alligators in three decades was permitted in 1995 (Rhodes, 1996). Texas: Within Texas, the American alligator ranges “from the Sabine River of East Texas to the Gulf of Mexico across the coastal marshes to the Rio Grande and west to around Interstate 35. This range includes about 120 counties with the highest concentrations occurring along the Gulf Coastal Plains” (TPWD, 2007b). It was estimated that there were around 290,000 alligators in Chambers, Jefferson, and Orange counties (TPWD, 2004; TPWD, 2007a) but no state-wide population estimates were reported to have been made (TPWD, 2004). In the Welder Wildlife Refuge, Texas, a density of 0.62 alligator/ha was found in 1998, representing an increase of 194% and 30% compared with 1973 and 1978 estimates respectively, from the same region (Altrichter & Sherman, 1999). A population index estimated that Lake Raven, a highly disturbed inland lake in Texas, supported 0.19 alligators per hectare, with most of these alligators being juveniles (Lutterschmidt & Wasko, 2006). The Texas Parks and Wildlife Department monitors the Texas alligator population via aerial nest surveys and night spotlight counts. Based on data acquired, TPWD allows egg collection, farming, and hunting activities for alligators by permit. Annual harvest records are kept, documenting harvest size, sex and county of take (TPWD, 2007c; TPWD, 2007d). The harvesting of alligators and their eggs is strictly regulated in Texas, through a system that includes, amongst other elements, hunting seasons, licenses, tags, hunting methods, and areas where hunting is permitted (TPWD, 2007c).

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During the period 1986-1991, between 950 and 2050 wild alligators were harvested per year in Texas. Harvest of farmed alligators in Texas began in 1989; 20 alligators were harvested that year, and two years later, in 1991, 700 alligators were harvested from Texas farms (Masser, 1993). A controlled alligator nuisance programme also exists in Texas (Johnson et al., undated; TPWD, 2007c). Alligators in Texas have been found to have a lower growth rate than that reported in other regions of the United States (16 cm per year compared with 20-30 cm per year in South Carolina, Florida, and Louisiana), and for this reason it has been suggested that more research is necessary to improve harvest management (Merendino et al., 2001). It has also been noted that in the coastal areas of Texas alligators may be somewhat limited to freshest sites along the salinity gradient, and as degradation of coastal marshes continues, most notably due to saltwater intrusion, alligators may be forced into more inland habitats (Merendino et al., 2001).

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USFWS (U.S. Department of the Interior Fish and Wildlife Service). 1987. Reclassification of the American Alligator to Threatened Due to Similarity of Appearance Throughout the Remainder of its Range. Federal Register, 52 (107): 21059-21064.

USFWS (U.S. Department of the Interior Fish and Wildlife Service). 1992. American Alligator (Alligator Mississippiensis). U.S. Fish and Wildlife Service, Tulsa, Oklahoma Ecological Services. Accessed April 2007. URL: http://www.fws.gov/southwest/es/oklahoma/gator.htm

USFWS (U.S. Department of the Interior Fish and Wildlife Service). 2007. Species profile: American alligator (Alligator mississippiensis). Accessed in April 2007. URL: http://ecos.fws.gov/speciesProfile/SpeciesReport.do;jsessionid=D3B860F0ED5D54AB3BB8E1FF519B8655?spcode=C000#conservationPlans

Woodward, A.R. 1987. Alligator Ranching Research in Florida, USA. In: Webb, G.J., Manolis, S.C. and Whitehead, P.J. (Eds). 1987. Wildlife Management: Crocodiles and Alligators. Surrey Beatty & Sons, Chipping Norton, NSW, Australia.

Woodward, A.R. 1996. Determination of appropriate harvest strategies for alligator management units. Final Report. Florida Game and Fresh Water Fish Commission, Tallahassee, Florida.

Woodward, A.R. 1998. Assessment of the economic viability of the Florida alligator industry. Final Report. Bureau of Wildlife Research, Division of Wildlife, Florida Game & Fresh Water Fish Commission, Tallahassee, Florida.

Woodward, A.R. & Cook, B.L. 2000. Nuisance-Alligator (Alligator mississippiensis) control in Florida, USA. P. 446-455. In: Crocodiles: Proceedings of the 15th Working Meeting of the Crocodile Specialist Group, IUCN, Gland, Switzerland.

Woodward, A.R. & David, D.N. 1993. The rise and fall of classic crocodilian skin prices: Where do we go from here? In: Crocodiles. Proc. of the Second Regional (Eastern Asia, Oceania, Australasia) Meeting of the Crocodile Specialist Group. IUCN-The World Conservation Union, Gland, Switzerland

Woodward, A.R. & David, D.N. 1994. Alligators. In: Hygnstrom, S.E., Timm, R.M. and Larson, G.E. (Eds.) Prevention and Control of Wildlife Damage. University of Nebraska.

Woodward, A.R. & Moore, C.T. 1990. Statewide Alligator Surveys. Final Report. Bureau of Wildlife Research, Florida Game and Fresh Water Fish Commission, Tallahassee, Florida. 24pp.

Woodward, A.R. & Moore, C.T. 1995. American Alligators in Florida. In: LaRoe, E.T., Farris, G.S., Puckett, C.E., Doran, P.D. & Mac, M.J. (Eds). 1995. Our living resources: a report to the nation on the distribution, abundance, and health of U.S. plants, animals and ecosystems. U.S. Department of the Interior, National Biological Service, Washington, D.C. 530 pp.

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REPTILIA: ALLIGATORIDAE

SPECIES: Caiman crocodilus crocodilus

SYNONYMS: - COMMON NAMES: South American Spectacled Caiman (English),

Caïman à lunettes (French), Swedish - sydamerikansk glasögonkajman;

RANGE STATES: Brazil, Colombia, Ecuador, French Guiana, Guyana,

Peru, Suriname, Trinidad and Tobago, Venezuela RANGE STATES UNDER REVIEW: All, with emphasis on the main exporters: Colombia,

Guyana, Surinam and Venezuela IUCN RED LIST: Not Evaluated (Caiman crocodilus: Lower risk/Least

Concern) PREVIOUS EC OPINIONS: Import suspensions for wild Caiman crocodilus

crocodilus were in place from 22/12/1997 until 24/09/2000 for Bolivia, Ecuador, Peru, and Trinidad and Tobago. The subspecies does not, however, occur in Bolivia.

At the species level, positive opinions were formed for Suriname on 22/07/1997 and for Ecuador, Peru and Trinidad and Tobago on 11/07/2000. Other opinions/suspensionsare in place at the species level in countries where the subspecies crocodilus does not occur, and are not discussed here.

TRADE PATTERNS: The Spectacled Caiman, Caiman crocodilus crocodilus, was selected for this review on the basis of high volumes of trade into the EU in 2004 (UNEP-WCMC, 2006). Owing to the extensive development of ventral osteoderms (boney inclusions), caiman belly skins are of inferior commercial quality compared to those of crocodiles and the American alligator, and usually only the lateral flank region is used. Because of the low value of the hide, caiman exploitation did not begin until the 1950s when stocks of the more valuable classic crocodiles became depleted (Ross, 1998; Britton, 2007). The value index of Caiman spp. skins was US$2/cm in the early 1980s, rising to US$9/cm in the late 1980s, and falling to 5 US$/cm in the early 1990s; between then and the year 2000, the value index has been between 5 and 7 US$/cm (MacGregor, 2002). Of the Caiman crocodilus crocodilus range States, the main exporters (and the principal modes of production) were: Venezuela (wild harvest and captive breeding), Guyana (wild harvest), Colombia (captive breeding), and Brazil (captive breeding, ranching under development) (Hutton et al., 2001). During the period 1996 – 2004, the reported trade in Caiman crocodilus crocodilus skins was between 26,000 and 56,000 per year (Caldwell, 2006).

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Germany (51%), Italy (39%) and Spain (10%) between them accounted for all EC imports of skins, and over 97% of EC-reported imports were exported by, or originated in, Venezuela (UNEP-WCMC, 2006). A summary of CITES trade data is provided in Tables 13-15. At the species level (Caiman crocodilus), EU Member States reported the import of 3 bodies (source I) from Brazil in 2001. Brazil did not report this trade. Additionally, Brazil reported the export to Spain of one skull (Source W) in 2003. Spain did not report this transaction. Also at the species level (Caiman crocodilus), EU-27 Member States reported, between 2001-2005, the direct import of two skulls and one carving (Source I) from Brazil, two skulls (Source I) from Ecuador, and one skull and one carving (Source I) from Peru. Exporting countries did not report these transactions. “Venezuela is the main supplier of Caiman c. crocodilus skins, almost all from wild-collected animals. Exports peaked in the late 1980s and early 1990s with quantities exceeding 100,000 skins in several years. More recently exports have gradually declined, possibly as a result of the farming of massive numbers of Caiman crocodilus fuscus in Colombia, but also because of high taxation of the caiman hunting industry” (Caldwell, 2006). “Guyana was the supplier of more than 350,000 skins between 1984 and 1989 but exports dwindled during the 1990s and there are no records of skins being exported between 1998 and 2000. Exports from Guyana resumed in 2000” and between around 2000 and 5000 live animals per year have been exported globally from Guyana since 2001 (Caldwell, 2006). Exports from captive-breeding operations in Brazil amounted to 6520 skins in 2000 but none subsequently (Caldwell, 2004). Colombia also exports small quantities (Caldwell, 2004). There have been undocumented but persistent allegations of extensive illegal trade in caiman skins, especially from Colombia, Brazil and Peru (Groombridge, 1982; Ross, 1998), with estimations that legal export figures reflected about one half of the true number of hides exported from Colombia and Peru (Groombridge, 1982). In the 1980s some authors pointed out that “the species [Caiman crocodilus] is hunted intensely for skins, […], the great majority being exported illegally from their countries of origin” (Luxmoore et al., 1988). More recently, illegal hunting and illegal trade were also considered to be key issues affecting Caiman crocodilus (Ross, 1998; Britton, 2007). In 2001, 2000 skins of Caiman crocodilus fuscus from Colombia were seized by the United States, and in 2002 Spain reported the seizure of 300 skins of Caiman crocodilus fuscus from Colombia (Caldwell, 2006). Table 12. CITES export quotas for Caiman c. crocodilus

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006

Guyana skins 20000 20000 20000 20000 20000 20000 20000 20000 20000 20000 live 10000 10000 10000 10000 10000 10000 10000 10000 10000 10000

Suriname live 980 925 925 925 925 925 925 925 925 925

Table 13. Main direct exports of Caiman c. crocodilus to EU-27, 2001-2005. Source W unless otherwise indicated.

Exporter Term Reported by 2001 2002 2003 2004 2005 Total

Guyana bodies Importer 4

(Source I) 10

(Source I) 14

Exporter

live Importer 2770 737 2134 1713 2815 10169

Exporter 2597 737 1312 1563 2630 8839

skins Importer 100 100

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Exporter Term Reported by 2001 2002 2003 2004 2005 Total

Exporter 100 100

Peru live Importer

Exporter 120

(Source C) 120

Suriname live Importer 50 14 3 67

Exporter 67 14 9 90

Venezuela large leather products Importer 7 7

Exporter

live Importer 150

(Source C) 150

Exporter

skin pieces (kg) Importer 60 60

Exporter 60 60

skins (sides) Importer 29957 27130 63272 100387 128296 349042

Exporter 31245 27130 63645 121111 243131

skins (unspecified units) Importer

1000 (Source C) 1000

Exporter 1000

(Source C) 1000

small leather products Importer 60 60

Exporter 300 10 60 370

tails Importer 600

(Source C) 600

Exporter

Importer 3550 3550

Exporter 4150 4150

Table 14. Main indirect imports of Caiman c. crocodilus to EU-27, 2001-2005.


Bolivia small leather products C Importer

Exporter 40 40

Colombia large leather products W Importer

Exporter 32 32

skin pieces C Importer

Exporter 1350 1350

skins (sides) C Importer 3534 3534

Exporter 10507 10507

skins (unspecified units) C Importer 2992 2992

Exporter

W Importer 315 315

Exporter 365 365

small leather products C Importer 3882 1739 863 6128 8055 20667

Exporter 4828 2395 934 596 8753

U Importer 2 1 5 8

Exporter 74 74

W Importer 15 1835 3 6 7 1866

Exporter 454 473 75 76 1078

tails C Importer

Exporter 2992 2992

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Guyana live W Importer 22 22

Exporter 22 22

skins (sides) C Importer 250 250

Exporter

W Importer 3000 3000

Exporter

skins (unspecified units) W Importer 500 15 515

Exporter 862 862

small leather products C Importer 8 1 50 59

Exporter 2 50 52

U Importer 1 1

Exporter 16 16

W Importer 237 2824 1897 86 5044

Exporter 1164 243 207 102 1716

Japan small leather products C Importer 20 20

Exporter

Nicaragua small leather products W Importer 50 50

Exporter

Sudan small leather products C Importer 1150 1150

Exporter

Unknown small leather products C Importer 55 6 61

Exporter

W Importer 33 33

Exporter 3825 3825

Venezuela garments, large leather products and shoes C Importer 75 35 0 4 0 114

Exporter 0 14 6 4 1 25

W Importer 773 95 33 158 162 1221

Exporter 545 26 113 70 224 978

skin pieces (kg) W Importer 185 120 114.5 200 619.5

Exporter

skin pieces (unspecified units) C Importer 680 680

Exporter

W Importer 247 575 1115 1937

Exporter 3111 71 3182

- Importer

Exporter 40 40

skins (sides) W Importer 483 2273 2446 6439 11641

Exporter 480 5 5643 1640 7768

skins (unspecified units) C Importer 2763 4767 7530

Exporter

W Importer 396 16 412

Exporter 45 2 47

small leather products C Importer 107 50 9 217 6 389

Exporter 513 57 5 12 0 587

U Importer 1 1

Exporter 76 76

W Importer 2662 2150 615 3970 3588 12985

Exporter 1819 2731 3236 4788 91 12665

tails (kg) W Importer

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Exporter 114.5 114.5

- Importer

Exporter 154 154

tails (sides) - Importer

Exporter 232 232

tails (unspecified units) W Importer 110 3 4748 2777 7638

Exporter 443 3878 4321

Table 15. Main direct exports of Caiman c. crocodilus to countries other than EU-25, 2001-2005. Source W unless otherwise indicated.


Colombia large leather products C Importer 19

Exporter

skins C Importer 5123

Exporter 19531

Guatemala live C Importer 600

Exporter

Guyana bodies W Importer 41

Exporter

large leather products W Importer 113

Exporter

live F Importer 170

Exporter

W Importer 8248

Exporter 9877

shoes W Importer 58

Exporter

skins W Importer 3915

Exporter 4216

Suriname live W Importer 107

Exporter 565

Thailand live C Importer 7

Exporter 13

Uruguay skins I Importer 26672

Exporter

Venezuela live R Importer 1300

Exporter 37652

W Importer 770

Exporter

shoes W Importer 101

Exporter

skin pieces (kg) W Importer 2448

Exporter 3411.7

skin pieces (m) W Importer

Exporter 12

skin pieces (unspecified units) W Importer 3972

Exporter 13

skins (sides) W Importer 18397

50


Exporter 17604

skins (unspecified units) - Importer 4236

Exporter

small leather products W Importer 111

Exporter 201

tails W Importer 2425

Exporter 37950

CONSERVATION STATUS IN RANGE STATES: Some authors recognize between three and five subspecies of Caiman crocodilus (e.g. King & Brazaitis, 1971; Groombridge, 1982; Luxmoore et al., 1988; Ross, 1998), whereas others have found no basis for the partition of C. crocodilus into subspecies (Busack & Pandya, 2001). The standard reference adopted by CITES, Wermuth & Mertens (1996), recognizes three subspecies Caiman crocodilus crocodilus, Caiman crocodilus fuscus and Caiman crocodilus apaporiensis and, for the most part, CITES trade data have been reported accordingly. The subspecies crocodilus and fuscus are listed in Appendix II of CITES, whereas the subspecies apaporiensis is listed in Appendix I. Taxonomic confusion, together with the difficulty in differentiating between subspecies, has resulted in a lack of literature at the subspecies level, especially for those subspecies that overlap in part of their ranges: Caiman c. crocodilus and Caiman c. fuscus. Moreover, there appeared to be evidence that the subspecies in trade were regularly misidentified (Luxmoore et al., 1988). Caiman crocodilus crocodilus occurs in northern parts of South America east of the Andes: Brazil, Colombia, Ecuador, French Guiana, Guyana, Surinam, Trinidad and Tobago and Venezuela (Groombridge, 1982; Luxmoore et al., 1988). Although the available information was patchy in many areas, Caiman crocodilus populations appeared to be doing relatively well in most countries (Ross, 1998), and surveys suggested that populations were in relatively good condition in most areas (Britton, 2007). Caiman crocodilus crocodilus was considered a highly adaptable caiman, quick to occupy available habitat, including that made by man (Groombridge, 1987). The species appears to have been quite resilient to commercial hunting for a number of reasons, particularly because they reproduce at a relatively small size, and hunting in many areas appears to have concentrated on the larger adult males (Rebelo & Magnusson, 1983; Herron, 1991; Mourao et al., 1996; Ross, 1998). The near extirpation of larger, sympatric species of crocodilians of greater commercial value means that Caiman now occupy habitats that were formerly dominated by Melanosuchus niger, Crocodylus intermedius and C. acutus (Magnusson, 1982; Ross, 1998). “Furthermore, in areas such as the llanos of Venezuela and Colombia and the Brazilian Pantanal, the proliferation of man-made water bodies (e.g. borrow-pits) has increased the carrying capacity for caiman populations in these habitats. Although they may be locally depleted, present populations may be larger than they were historically. […] Local extirpation adjacent to urban areas and intensive agriculture is observed, but wherever their habitat remains intact, this adaptable small crocodilian appears to remain abundant” (Ross, 1998). Where harvest regulations and use programmes were enforced on a sustainable basis, populations were reported to be stable or increasing (Ross, 1998). Most sustainable use programmes rely upon regular cropping of wild populations, the long-term effects of which need to be investigated. Farming and ranching programmes, while they exist, may be uneconomical in the long run, given the comparatively low value of the hide and the number of

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animals which need to be culled in order to produce a profitable amount of hide (Britton, 2007). In fact, at the workshop “Rational Use of Crocodiles and Caimans” held in Peru in 1999, it was recommended for Caiman crocodilus that “Neither breeding in captivity nor their management for ranching is recommended since both practises are very expensive, taking into account the low price in the international market. It was recommended that their population status be evaluated for the development of a program of direct cropping (hunt), following the example of the Venezuelan program” (Larriera, 1999). Reported to be mainly threatened by illegal hunting for skins and lack of enforcement to curtail smuggling (Groombridge, 1987). In the late 1960s, most wild Caiman crocodilus crocodilus populations were considered to be declining due to slaughter by hide hunters and capture by live animal collectors (King & Brazaitis, 1971). In the 1980s, populations of Caiman c. crocodilus appeared to be at least stable throughout much of South America, with the exception of Colombia, Peru and some parts of Guyana, and it was considered that it would be considered non-threatened instead of vulnerable, if it were not for the reported increase in hunting pressure, and the lack of data – due to trans-border smuggling – concerning the extent of exploitation in particular areas (Groombridge, 1982; Groombridge, 1987; Luxmoore et al., 1988). In fact, illegal hide hunting, illegal trade and lack of adequate enforcement to control smuggling seemed to be the major factors contributing to the pressures placed upon the subspecies (Groombridge, 1982; Groombridge, 1987; Luxmoore et al., 1988; Ross, 1998; Britton, 2007). Improved enforcement and CITES implementation appeared to have reduced the flow of illegal skins (Ross, 1998).

BRAZIL: Caiman crocodilus crocodilus occurs throughout the Amazon basin of Brazil with the exception of a few southern tributaries (King & Brazaitis, 1971; Luxmoore et al., 1988). It was reported to occur in the following states: Amazonas, Para, Rondonia, Maranhao, Amapa, Acre, Roraima, Tocantins, Mato Grosso, Goias, and Brasilia (Pinto, 2002). With an estimated population of over 1,000,000 individuals, the spectacled caiman was considered the most common of the Brazilian crocodilians, although some populations are locally reduced (IBAMA-RAN, 2007). A study, conducted between 1991-1992 in the Anavilhanas Archipelago, located in the lower Rio Negro, Central Amazonia, found 466 Caiman c. crocodilus in 81 survey sites along 82 km of bank. Of these, 88 were hatchlings. Densities varied from 0 to 58 per km of shoreline in lakes, and from 0 to 7 per km of shoreline in canals. Zero counts occurred in 10 of the 81 survey sections, but the species was present around the survey sections in all 81 sites (Da Silveira et al., 1997). Another study conducted between 1993-1996 in Jau National Park, Brazilian Amazon, found Caiman crocodilus densities of between 0 and 2 caimans/km in a total of 30 sites; densities in most sites were between 0.2 and 0.9 caimans/km (Rebelo & Lugli, 2001). During surveys of six sectors within the Mamiraua reserve in Brazil in 1995, 4586 Caiman crocodilus were counted, 24% of them with a snout-vent length larger than 70 cm (Da Silveira & Thorbjarnarson, 1998). During a survey along Rio Purus (Amazonas state, Brazil) Caiman c. crocodilus was observed in all the 20 sites (140 km) that were sampled at night (Da Silveira, 2001). Another survey along 280 km of shoreline in 22 water bodies within the Piagacu-Purus Sustainable Development Reserve found around 3000 C. crocodilus; additionally, 28 nests of C. crocodilus were found in 11 out of 15 localities searched, and local people poached eggs in 5 of them (Von Muhlen et al., 2006). A 2004-2005 survey in the Javaes River, Araguaia National Park (Tocantins state, Brazil) found C. crocodilus densities of 17.9 (2004 dry season), 0.28 (2005 wet season) and 10.5 (2005 dry season), with 10% of the C. crocodilus being in the <50cm size class, 37% in the 50-140 cm size class, 38% in the 140-180 cm size class and 15% in the >180 cm size class (Malvasio & Salera, 2006).

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It was reported that “Brazilian crocodiles are threatened by gold mining and related mercury- and lead-contamination of animals and their habitat, deforestation, increasing use of land for agriculture and destructive forestry, increased human incursions, and commercial and subsistence hunting for skins and meat” (Brazaitis et al., 1996). In 1967, hunting of all wildlife, including caiman, was banned in Brazil –Law 5.197/67-, captive breeding and ranching remained the only possibilities for commercial use of the species (IBAMA, 2003). Caiman c. crocodilus reportedly recovered rapidly since the ban, and around a decade later it was caught frequently in gill nets in floodplain lakes, even near Manaus, and could be seen without difficulty on tributaries of the Amazon such as the Rio Negro and Rio Tocantins (Groombridge, 1982; Luxmoore et al., 1988). It was also reported to occur in high numbers in most areas, and populations appeared to be limited by habitat preference rather than by disturbance by man, having even colonised the artificial lakes formed behind the Transamazonian Highway, and having increased in areas where populations of Melanosuchus niger had been reduced by hunting (Magnusson, 1982). Because the size and remoteness of the Amazon basin made it very difficult to enforce the 1967 ban, illegal caiman skin hunting continued at least through the 1970s (Magnusson, 1982; Rebelo & Magnusson, 1983; Luxmoore et al., 1988), largely driven by overseas demand, particularly from France, Germany, Italy and the US (Luxmoore et al., 1988). Nonetheless Caiman crocodilus remained common in many areas of Brazil, even under considerable hunting pressure (Rebelo & Magnusson, 1983). During the 1980s and 1990s, no evidence of illegal caiman skin trade was found in the Brazilian Amazon, except for 55 skins confiscated in 1998, over half of them belonging to females or subadult males (Da Silveira et al., 1998). Following the ban the market for skins disappeared and hunters began selling caiman meat instead. By the early 1980s a trade in salted meat from the Amazonas to Para state in Brazil and to Colombia was established (Da Silveira & Thorbjarnarson, 1998). Based on data collected in 1995, it was estimated that 2865 Caiman crocodilus were illegally hunted for their meat per year in the Mamiraua Reserve in Brazil; 97% of these were sexually mature, and the sex ratio (12:1) was strongly male biased. However, this harvest appeared to be less detrimental to the caiman populations than the harvest for skins. This was mainly because the meat was less valuable than skins and therefore small individuals and areas of difficult access were left mostly untouched, as they were not worth the hunters’ time (Da Silveira & Thorbjarnarson, 1998). A survey in Rio Purus (Amazonas state, Brazil) found through interviews with caiman hunters that illegal exploitation of caimans for meat happens throughout the whole year in the area (Da Silveria, 2001). It also stated that “the information gathered suggests, as a first approximation, that caiman populations in the area must be huge and, as stated by Bill Magnusson, no doubt it is manageable to hunt thousands of caimans per year in this area with such a complexity and abundance of habitats” (Da Silveria, 2001). A further study, however, based on data from 2004-2006, estimated that annually, 3343 Caiman crocodilus were taken illegally for meat in the Piagacus-Purus Sustainable Development Reserve (Amazonas state, Brazil), particularly in the northern Cuiuana region and between the Reserve and the Amazonas River, mostly to sell this meat in the Para state; the study suggested that “this exploitation is unsustainable and a ban of at least five years is necessary for a future caiman management program in the Piagacu-Purus Sustainable Development Reserve” (Marioni et al., 2006).

COLOMBIA: In Colombia, Caiman crocodilus crocodilus occurs to the east of the Andes, in the Amazon and Orinoco drainages (Luxmoore et al., 1988; Ross, 1998). Few recent data on the status of Caiman crocodilus crocodilus in Colombia was available. A census of crocodilians in Colombia carried out in 1994-1995 (Barahona et al., 1996) found a total of 2133 Caiman crocodilus crocodilus in 54,469 ha of aquatic surface (see Table 16). A census carried out in 1995-1997 indicated that although some populations had decreased or even

53

disappeared, the status of the species in the country as a whole could be considered as “Lower Risk” (Rodriguez Melo, 2002). Table 16: Total number and density of Caiman crocodilus crocodilus at different locations within the Orinoco and Amazon hydrographical areas of Colombia, 1994-1995. Source: Barahona et al., 1996.

Sampling location Observed individuals

Density (caimans/ha)

Rio Ele-Cravo Norte 627 3.82

Rio Guaviare-Inirida 506 0.17

Rio Inirida 279 0.17

Rio Putumayo 82 0.03

Rio Curilla 30 0.22

Rio Caucara 32 0.12

Quebrada La Playa 14 0.14

Laguna La Playa 14 0.1

Cocha Yarinas 2 0.03

Esteracocha 21 0.13

Sunichocha 93 0.5

Rio Caqueta 147 0.45

Rio Bumelo 20 0.22

Rio Caguan 25 0.4

Sampling location Observed individuals

Density (caimans/ha)

Lago El Mosco 91 2.6

Rio Amazonas 9 0.01

Quebrada Pichuna 3 0.05

Quebrada Matamata 5 0.09

Rio Atacuari 7 0.12

Rio Loretayacu 22 0.04

Rio Amacayacu 3 0.01

Rio Borahuazu 12 0.05

El Soco 32 1.03

Islas Colombia 24 0.51

Pozo Cacharama 14 1.27

Lago Tarapoto 6 0.06

Lago Garzachocha 13 0.4

TOTAL 2133

In the Casanare river basin, the size class distribution was 38% class IV, 32% class III and 30% class II, which was interpreted as an indication that the population had not been subject to illegal hunting (Barahona et al., 1996). However, size class distributions in the other river basins, with small proportions of class IV individuals, indicated that these populations had been exploited (Barahona et al., 1996). The subspecies was reported in the late 1970s to be rarely hunted on the Vaupes and Guayabero-Guaviare rivers because the price of petrol, salt and air transport to Villavicencio and Bogota were too high (Luxmoore et al., 1988). Juveniles were seen often in the 1980s, with some adults reappearing. There was a healthy population in the “Tomo-Tuparro” Faunistic Territory, a 600 000 ha reserve maintained by INDERENA (Luxmoore et al., 1988). In 1975, 350 caimans, mostly juveniles were counted in the lagoons and backwaters of the Capanaparo River (Medem, undated, in Luxmoore et al., 1988). However, some hunting continued in the Llanos (Orinoco plains), and the fact that most hides were between 30-60 cm may suggest that adults had been virtually wiped out from that area (Medem, 1980).

Although the export of caiman skins under 1.5 m was prohibited in Colombia in 1973, a trade in this species’ skins under the minimum legal length was reported to have continued at least until the 1980s (Luxmoore et al., 1988). Also from 1973, only captive bred Caiman c. crocodilus skins were permitted for export (Ulloa Delgado, 1996). Despite the restrictions, Luxmoore et al. (1988) reported that: “many skins came from individuals killed in Peru and Brazil and smuggled over the border. The minimum legal size restrictions are not enforced. Along the Caribbean coast between Cienaga Grande and the Magdalena River illegal hunting is common, also on the lower and middle Magdalena. Almost all hides are of hatchlings or juveniles. Some refuge is found in lagoons along the lower and middle Magdalena Valley, where aquatic vegetation has become more extensive with the decline of manatees. Additionally, hatchlings are killed in large numbers and are preserved as curios for sale to tourists. For many years, C. crocodilus hatchlings from Colombia and Ecuador had provided crocodilian pets for the world pet trade. It is estimated that legal export figures reflect about one half of the true number of hides exported from Colombia and Peru, a large number of those animals killed spoil before reaching the tanneries. In many instances, the animal is killed for its flanks only, the reminder of the body

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left to waste. Marked habitat destruction has occurred in Colombia including the officially protected mangroves of the Isla de Salamanca National Park”. More recently, allegations of illegally harvested wild caiman skins sold to farms and tanneries, where they are mixed with legal farmed skins have been made in the last few years, although they have not been proven (Ross, 2003; MacGregor, 2006). Because of these concerns, the Colombian CITES Management Authority put in place domestic restrictions on the size of skins for export, with a self-imposed maximum C. crocodilus skin size limit of 125 cm (CITES Notification to the Parties No. 2002/031). Moreover, any venture breeding crocodilians in Colombia is required to operate non-commercially for an experimental phase (a minimum of two years) before the CITES Management Authority issues a commercial licence. In addition, the Ministry of the Environment obliges all farms to provide crocodilians for restocking wild populations (MacGregor, 2006). Since the 1980s, there has been a strong investment in crocodilian captive breeding operations in Colombia, and in the early 1990s Colombia was the largest producer of crocodilians in the world, a position that has been maintained to date (MacGregor, 2006). The great majority of caiman skin exports from Colombia are from captive-bred sources, and the subspecies most farmed and traded is Caiman crocodilus fuscus (Ulloa Delgado, 1996; MacGregor, 2006).

ECUADOR: Found in the Amazonian region of the country (Luxmoore et al., 1988). Larriera et al. (2005) stated that in Ecuador C. crocodilus was abundant in the wild, that there was a subsistence market for their meat, but there was no management plan for the species. In 1973, the subspecies was not considered to be seriously endangered (Medem, 1973, cited in Groombridge, 1982). In 1983, populations were said to be thriving in some areas, and there were reports that they had increased in recent years (Asanza, 1984), despite the illegal movement of skins by hunters and traders over the border into Peru and Bolivia at least up to the 1980s, when border controls were improved and the trade declined (Groombridge, 1988). In 1994, crocodilians on the Curaray River, Ecuador, were surveyed: C. crocodilus was observed in 10 of 13 lagoons and on all river survey routes (Hines & Wilkinson, 1996). The mean densities found for the species were 3.85 caimans/km in lagoons and 1.73 caimans/km on river routes. The size class distribution was: 47% <0.6m, 32% 0.6 - 1.2m, and 20% >1.2m (Hines & Wilkinson, 1996). A study on the growth rates of Caiman c. crocodilus in the Ecuadorian Amazon suggested that hunting pressure on crocodiles may benefit the species by reducing interspecific competition. The lower growth rate of Melanosuchus niger means that it has a more reduced capacity to recruit reproductive individuals (Vallejo et al., 1996). For example, a survey carried out in six lakes in the northern part of the Ecuadorian Amazon during 1992-1994, found that the C. crocodilus/M. niger ratio was related to the accessibility by humans to each of the lakes, with less accessibility accounting for less hunting pressure, which benefited M. niger (Ron, 1996). The C. crocodilus densities found during this study ranged between 35 and 44 caimans/km of shoreline (Ron, 1996). Another study carried out during the same period estimated a total number of 15 C. crocodilus at Zancudococha Lake, north-eastern Ecuadorian Amazon, with a density of 0.29 ind/km (Vallejo, 1996).

FRENCH GUIANA: Reported in 1973 not to be in serious danger of extinction (Medem, 1973, in Groombridge, 1982). However, no recent information on the status of the species in French Guiana was identified. A number of general genetic studies have been published in the last few years that considered C. crocodilus in the country (Farias et al., 2004; Vasconcelos et al., 2006). No trade in Caiman crocodilus crocodilus has been reported since at least 1996 from French Guiana (Caldwell, 2006).

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Caiman crocodilus was listed under Article 3 of Ministerial Decree of 15 May 1986, which prohibited their use, taxidermy, purchase or sale, and which allowed their transport only if they had been legally acquired outside the territory (Luxmoore et al., 1988). However, Farias et al. (2004) reported that C. crocodilus was not protected by law at present in French Guiana, although poaching seemed to be currently limited by the fact that most of its range in the country is within a nature reserve system.

GUYANA: In 1973 populations in Guyana were reported to be abundant in most areas but declining in areas where habitat had been lost to agriculture, particularly rice fields (Medem, 1973, in Groombridge, 1982). A survey in the mid-1990s of Caiman crocodilus crocodilus at the Iwokrama Reserve in central Guyana during the wet season indicated that this population had the same demographic structure of that of a healthy, unhunted population (Alstair, 1996). However, no recent information on the status of the species in Guyana was identified All exports were prohibited in 1987, but later that year the ban was lifted and an export quota system was introduced (Luxmoore et al., 1988). During the last decade (1997-2007), this quota has been set at 20,000 skins and 10,000 live individuals per year.

Guyana has in place a harvesting programme for wild C. crocodilus (Britton, 2007) and is with Colombia the second largest exporter of the species, although exports from these two countries are very reduced compared to those from Venezuela (Caldwell, 2006).

PERU: Found in the Amazonian region (Groombridge, 1982). Larriera et al. (2005) reported that there was no recent studies carried out on this species in Peru; however, its status in the country was considered to be “good” (Larriera et al., 2005). Reported to be seriously depleted in 1973, on the basis of a rapid decline in numbers of hides exported (Medem, 1973, in Groombridge, 1982). It was reported to be in decline two decades later (1999) in the department of San Martin, northeast Peru (Schulte, 1999). In the Manu National Park it was said to be frequently observed in 1979 (Pereya, 1979, in Luxmoore et al., 1988). Several Caiman crocodilus were seen in Chira River, north-western Peru, which is not the natural range of the species; this was considered to be the result of an unofficial introduction (Trelancia, 2006). No trade in this species was permitted under Decree No. 934-73-AG of 1973 (Luxmoore et al., 1988), until 1999 (INRENA, 2000).

SURINAME: Found throughout the northern plains and swamps (Glastra, 1983). Reported in 1973 to be abundant in suitable habitat in Suriname, with populations as close to an untouched state as could be found (Medem, 1973, in Luxmoore et al., 1988). By 1983, the species was still locally common along the north coast and its status was not thought to be critical (Glastra, 1983). Large-scale hide hunting was reported in two consecutive dry seasons in 1978 and 1979 in the Coesewijne river and swamp area, causing severe losses estimated at 1500-2000 caimans (Glastra, 1983). A survey was conducted in 1981 in this area, estimating dry-season densities of around 6 caimans/km, and the total population at 1000-1500, which implied that at least half of the population had been killed in 1978 and 1979 (Glastra, 1983). The area was surveyed again during 1982-1984 (Ouboter & Nanhoe, 1988; Ouboter & Nanhoe, 1989), when an average density of 35 caimans/km and a total population of 4700 individuals was estimated. The size distribution, consisting of a large proportion of juveniles and few large adults, was considered to be that of a recovering population (Ouboter & Nanhoe, 1989). Suriname has not been an exporter of Caiman crocodilus crocodilus skins in recent years, but it exports some live wild-caught animals for the pet industry (Caldwell, 2006). Export quotas have

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been set at 925 live individuals per year since 1998 (UNEP-WCMC, 2007), but reported exports have been considerably lower than this figure (Caldwell, 2006).

TRINIDAD AND TOBAGO: Known to occur in both Trinidad and Tobago (Murphy, 1997). Reported to be abundant in parts of Trinidad including in the vicinity of the capital (Medem, 1973, in Groombridge, 1982). In Tobago it was formerly widespread throughout most of the rivers and marshes but its range was by the 1980s limited to a few rivers along the South coast. It was reported to be quite abundant in the lake behind the Hillsborough Dam (Hardy, 1982, in Luxmoore et al., 1988). The species was protected in Trinidad in 1958, yet in some parts it was still illegally hunted on a limited scale at least up to the 1970s; the population of Tobago was not subject to hunting for skins (Medem, 1973, in Groombridge, 1982; Luxmoore et al., 1988).

VENEZUELA: Caiman c. crocodilus is found throughout the Llanos and the Guyanan region as far as Amazonia. Also recorded from numerous localities along the north coast. The north-west of the country is populated by C. crocodilus fuscus but the exact range of the two subspecies was not known. However it was thought that the boundary is the Yaracuy River, with only C. crocodilus crocodilus eastwards from there (Gorzula, 1987; Luxmoore et al., 1988). C. c. crocodilus was not listed in the Venezuelan Red Data book (Rodriguez and Rojas-Suarez, 1995). In response to degradation of caiman populations from commercial hide hunting, in 1973 hunting was made illegal and populations subsequently increased so that in the late 1970s, the subspecies (C. c. crocodilus) was considered to be relatively abundant in Venezuela. (Groombridge, 1982; Gorzula, 1987; Luxmoore et al, 1988). It was reported that: “in Estado Bolivar it is difficult to find a body of water where this species does not occur; the population in this state may be as large as half a million. However, they are so easily hunted that they could be quickly eliminated from many areas. In very dry years populations in savannas may be reduced by as much as 80% because of mortality in small individuals. However, populations seem able to recuperate. In the last 20 years the construction of dams and reservoirs on cattle ranches has greatly increased available habitat” (Gorzula, 1981, in Groombridge, 1982). Up until the 1980s, no long-term monitoring of crocodilian populations had been carried out in Venezuela, and about 72% of the country lacked records of any crocodilian species (Gorzula, 1987). Thorbjarnarson & Velasco (1999) suggested that monitoring of Venezuelan caiman populations in the late 1990s had been sporadic, and it had lacked clearly defined objectives and standardization. A census in 1982-1983 estimated that in 39 private ranches in the states of Apure, Guarico and Bolivar, the total population was 700,000 to 1,000,000, from which a cull of 70,000 was authorised. In the Guyana region, the density of C. crocodilus was estimated to be 23.4 per km in lakes and ponds, and 2.5 per km in rivers. It was not thought that the populations of caiman in this region were under threat (Gorzula and Paolillo, 1986). The density of C. crocodilus crocodilus in the state of Anzoategui was estimated in 1984 to be 26.3 per km in the Rio Unare and 1.2 per km in the Rio Neveri. It was thought that the populations were not only abundant, but may have increased. Reasons for this included the creation of new habitat in the form of reservoirs, and the elimination of Crocodylus acutus, a potential competitor. Adverse ecological factors such as saline conditions seemed to be the main limitation to the spread of C. crocodilus populations (Seijas, 1986).

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A Caiman crocodilus nest survey carried out in 1994 across different ecological regions in Venezuela found 1905 nests in 242,404 ha (Velasco et al., 1996),. The highest density (0.23 nests/ha) was found in Cajon de Arauca, and the lowest (0.03 nests/ha) in Alto Apure (Velasco et al., 1996). Another survey in the mid-1990s found 93,944 Caiman crocodilus in 514,096 ha of Venezuelan floodable plains across six ecological regions, with a terrestrial density of 0.18 individuals/ha (Villarroel et al., 2004). A Caiman crocodilus population evaluation carried out in 1992 in the Orinoco delta found a density of 0.91 ind/ha during periods of high level water, and the following size class structure: 41% class II, 41% class III, and 18% class IV. During periods of low water, the density was 2.61 ind/ha, and the size class structure was: 35% class II, 46% class III, and 18% class IV (Velasco & Blanco, 1996). It was concluded from this evaluation that between 10% - 20% of the class IV individuals (i.e. 12,500 – 25,000 individuals) could be harvested in a sustainable manner in the Orinoco Delta (Velasco & Blanco, 1996). In 1995, a census was carried out to assess the status of Caiman crocodilus in the ecological region of Guarico, where hunting had been banned for 4 years (since 1992) due to the low proportion of adult males that was then found. In 251,115 ha, 25,859 individuals were found: 33% class II (subadults), 43% class III (adults), and 24% class IV (adult males), with a density of 0.10 caimans/ha (Colomine et al., 1996). This density was slightly lower than the 0.13 caimans/ha found in 1992, reflecting the slow recovery rate of these populations since the hunting ban was imposed. The study also pointed out the existence of a strong illegal hunting pressure on eggs and juveniles in the region, which would reduce the recruitment rate in wild populations (Colomine et al., 1996). A survey of Caiman crocodilus carried out in 2000 in the Hoya de Arismendi ecological region found a density of 0.11 caimans/ha, and the following size class distribution: 21% class IV , 60% class III, and 19% class II (Velasco & De Sola, 2000b). These results were similar to those obtained during a 1996 survey, which was considered to demonstrate that the programme of commercial use of Caiman crocodilus in the flooded llanos of Venezuela didn’t affect the wild populations of this species (Velasco & De Sola, 2000b). In 2002, Villarroel et al. (2002) assessed the status of Caiman crocodilus populations in Guarico and Llanos Boscosos ecological regions, finding densities of 0.05 caimans/ha and 0.01 caimans/ha, respectively, for each of the regions. These values were much lower than those found in previous censuses: 0.13 in 1992 and 0.09 in 1995 for Guarico region, and 0.15 in 1992 and 0.12 in 1996 for Llanos Boscosos region. This decrease in density was interpreted to be a result of the extreme drought conditions present during the period of study. The reduction in population density was therefore not thought to have been caused by over-exploitation or illegal hunting; however, suspension of licenses and strict surveillance of illegal hunting was recommended (Villarroel et al., 2002). The 2002 census revealed the following size classes: 17% size class II, 50% size class III, and 33% size class IV. This represented an increase in the proportion of classes III and IV and a reduction in the proportion of class II compared to the 1992 and 1996 censuses (Villarroel et al., 2002). Venezuela is the main exporter of Caiman crocodilus crocodilus (Caldwell, 2006). Regulated harvesting of wild Caiman crocodilus crocodilus has been permitted in Venezuela since 1983, under a programme directed by the Ministry of the Environment (MARN) allowing harvesting on authorised private ranches in the Venezuelan Plains (and between 1993 and 1997 also in the Orinoco Delta), of a proportion of the size class greater than 1.80 m, which are all adult males (Velasco, 2002; Velasco et al., 2003; Britton, 2007). Two years after the harvest programme started, in 1985, a one-year ban was imposed by a Ministerial Resolution, reportedly due to abuse and mismanagement in issuing hunting tags (Luxmoore et al., 1988). Following this ban, the programme entered a period of moderately stable harvests (85,000-165,000 caimans annually) between 1987 and 1991, followed by a great

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reduction in harvest in 1992 after a drop in skin prices. The number harvested subsequently increased slowly until another ban on hunting was imposed in 1996 due to “an overstock” of skins in Venezuela and the desire to census caiman populations without the interference from hunting activities. The total number of caiman legally harvested from 1983 through 1995 was 1,043,874 (Thorbjarnarson & Velasco, 1999). Harvest was re-established in 1997 and 126,778 caimans were harvested during the period 1997-2001. The decrease in the amount of harvested individuals since 1998 was attributed to the South Asia economic crisis (Velasco et al., 2002) and to low prices, low demand and increasing government taxes, which made harvesting caimans unprofitable for landowners (Velasco & De Sola, 2000a). The Caiman crocodilus harvest quotas were initially (in 1983) established based on estimates of the caiman population for each ranch requesting a permit, the quotas being 8-15% of the total population. Beginning in 1988 the harvest quota was based only on the number of caiman larger than 1.8 m total length; approximately 20% to 25% of this population subset was legally harvestable. These harvest quotas were generally viewed as conservative and likely to result in a sustainable harvest (Thorbjarnarson & Velasco, 1999). Since 1993, the harvest quota for caiman was calculated from average values of caiman density and population size-class structure in each of seven ecological regions within the Venezuelan Llanos (De Sola & Velasco, 1995; Thorbjarnarson & Velasco, 1999). These seven are: Llanos Boscosos, Hoya de Arismendi, Guarico, Bajo Apure, Aguas Claras, Cajon de Arauca, and Alto Apure (Velasco et al., 2003). Between 1988-2000, 75% of harvested skins in Venezuela were above 2.4 m of total length, well over the minimum size required, which was interpreted as a positive result in terms of the impacts of harvest on wild populations (Velasco et al., 2002). Velasco (1996) had also previously stated that survey results indicated that the Venezuelan program of exploitation of natural populations of caiman in the flooded llanos was sustainable. Velasco et al. (2003) compiled information on Caiman c. crocodilus for five of these ecological regions in 1996 and for the Bajo Apure region in 1999 (see Table 17). Survey data was collected over 532,447 ha for harvested (53% of the area surveyed) and non-harvested (47% of the area surveyed) populations within each of these regions, in order to assess the effects of harvest on caiman densities and population size structures (Velasco et al., 2003). The density of caimans was higher in the areas which had been subject to repeated harvest than in those which had remained unharvested. This was partly related to the fact that since the harvesting programme is voluntary, it was more likely to be implemented on those ranches where caiman were more abundant in the first place (Velasco et al., 2003). Similarly, the proportion of individuals over 180 cm was consistently higher in the harvested areas (>20%) than in the non-harvested areas (<20%) (Velasco et al., 2003). Table 17: Caimans/ha in each of the six ecological regions in the Venezuelan Llanos, with and without harvest. Source: Velasco et al. (2003).

Alto Apure Bajo Apure Aguas Claras

Cajon de Arauca

Llanos Boscosos

Hoya de Arismendi

With harvest 0.3 0.21 0.17 0.05 0.17 0.25

Without harvest 0.2 0.17 0.13 0.05 0.07 0.1

Velasco et al. (2003) concluded from these results that “there is no indication that populations have declined significantly as a consequence of harvesting, nor that adult males, which are selectively harvested, have become rare. The program brings considerable economic benefits to the landowners involved, who now have invested interest in caring for both caimans and their habitats. […] That densities were generally higher in harvested areas than in non-harvest areas could reflect hunting biases to some degree, but does allow rejection of some of the more dire predictions about the likely impacts of harvest. Caiman populations subject to this type of harvest (class IV males) still exist in high density despite 20 years of harvest. […] Monitoring at this level of resolution should be adequate for ensuring that the Venezuelan

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caiman harvest is sustainable, and for giving early warning of any problems, be they related to harvest or not. The monitoring program is thus sufficient to allow compliance with the purpose of Article IV.2b of CITES, which obligates the exporting State to provide assurance that the harvest for international trade is not detrimental to the survival of that species”.

Illegal commercial hunting of crocodilians, principally of Caiman crocodilus, was known to occur in Venezuela in the 1980s in the southwestern Llanos and Orinoco Delta, from where skins were exported, as contraband, to Colombia, Guyana and Trinidad (Gorzula, 1987). Illegal hunting also occurred in the 1990s (Rodriguez, 2000) and was considered to be one of the most problematic aspects of the harvest programme (Thorbjarnarson & Velasco, 1999). This primarily involved hunting caimans illegally in a variety of locations to sell them to landowners with legal hunting permits. These landowners did not want to reduce the size of the caiman population on their property because they thought it might jeopardize or reduce the size of future harvests. Therefore these infractions did not necessarily increase the total number of caiman killed, but resulted in an overharvest in some regions and a reduced number of killed caimans on many of the ranches with hunting permits (Thorbjarnarson & Velasco, 1999). Habitat modification, especially for cattle ranching, appears to have had an overall positive effect on populations of Caiman crocodilus (Gorzula, 1981, in Groombridge, 1982; Gorzula, 1987; Thorbjarnarson & Velasco, 1999).

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Pinto, M.G.M. 2002. Quelonios e Jacares do Brasil. IBAMA-RAN (Instituto Brasileiro de Meio Ambiente – Centro de Conservacao e Manejo de Repteis e Anfibios). URL: http://www.ibama.gov.br/projetos_centros/centros/ran/quelonios_Jacares.htm, Accessed 16th May 2007.

Rebelo, G.H. & Lugli, L. 2001. Distribution and abundance of four caiman species (Crocodylia: Alligatoridae) in Jau National Park, Amazonas, Brazil. Revista de Biologia Tropical. 49 (3-4).

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REPTILIA VARANIDAE

SPECIES: Varanus salvator

SYNONYMS: Monitor bivittatus celebensis, Monitor exilis, Stellio

salvator, Varanus vittatus COMMON NAMES: Common Water Monitor (English), Malayan Monitor

(English), No-mark Lizard (English), Plain Lizard (English), Rice Lizard (English), Ring Lizard (English), Two-banded Monitor (English), Water Monitor (English), Varan à deux bandes (French), Varan aquatique commun (French), Varano de dos bandas (Spanish), bandvaran (Swedish)

RANGE STATES: Bangladesh, Brunei Darussalam, Cambodia, China,

Hong Kong, China, India, Indonesia, Lao People's Democratic Republic, Malaysia, Myanmar, Philippines, Singapore, Sri Lanka, Thailand, Viet Nam

RANGE STATES UNDER REVIEW: All, with emphasis on Indonesia and Malaysia, the

main exporters IUCN RED LIST: Not Evaluated PREVIOUS EC OPINIONS: Imports from China, India and Singapore are

currently suspended (since 22/12/1997). Current positive opinion for Indonesia (formed 23/06/1999) and Malaysia (formed 05/09/2002). Imports were suspended from Bangladesh, Brunei Darussalam,

Cambodia, Myanmar, Philippines and Viet Nam from 22/12/1997 until 24/09/2000.

TRADE PATTERNS: Varanus salvator was selected for this review following a high volume of trade in wild sourced skins into the European Community in 2004 (UNEP-WCMC, 2006). Table 18. CITES export quotas for Varanus salvator

1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007

Indonesia

432000 5400 live; 408600 raw skins

450000 raw skins + live

471200 live and skins

5400 live; 444600 skins

5400 live; 444600 skins

6000 live; 444000 skins

5400 live; 446600 skins

5400 live; 444600 skins

9000 live; 441000 skins

6000 live; 441000 skins and skin products

Malaysia

12000 live; 160000 skins

Malaysia and Indonesia were the main exporters of Varanus salvator, with the remaining exports from India, the Philippines and Thailand (see Tables 19-21). Exports from the United States did

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not generally involve wild-caught animals. Exports from Singapore were re-exports of animals originating elsewhere. Table 19. Main direct exports of Varanus salvator to EU-27, 2001-2005. Purpose P unless otherwise indicated.

Exporter Importer Term Source Reported by 2001 2002 2003 2004 2005 Total

Indonesia Austria live W Importer

Exporter 45 45

Belgium live W Importer 10 10

Exporter 24 176 200

small leather products - Importer

51 (unspecified purpose) 51

Exporter

Czech Rep. live W Importer 7 (Purpose

B) 7

Exporter

Importer 40 40


France live F Importer 39 188 227

Exporter 58 89 108 255

W Importer 86 98 152 336

Exporter 80 78 72 30 260

skins W Importer 277 277

Exporter 3 350 353

small leather products W Importer

Exporter 2 2

Germany live W Importer 40 35 129 334 234 772

Exporter 112 85 197 818 268 1480

skins W Importer 1000 1000 2000

Exporter 1000 1000 2000


Exporter 6 6

Greece live W Importer 6 10 16

Exporter 6 10 16

Hungary live W Importer

Exporter 10 10

Italy live W Importer

Exporter 60 60

skins F Importer

Exporter 500 500

W Importer 38700 10000 3500 25000 61515 138715

Exporter 39700 16000 11500 28000 46517 141717


Exporter 20 8 68 96

Latvia live W Importer

Exporter 5 5

Malta live W Importer

66


Exporter 100 100

Netherlands live W Importer 16 24 51 91

Exporter 35 34 41 10 120

Spain live W Importer 25 54 39 15 133

Exporter 30 105 89 101 27 352

skins W Importer 2200 7100 6000 7100 15000 37400

Exporter 8000 5600 6000 9700 15000 44300


Exporter 8 8

Sweden live F Importer 4 4

Exporter 5 5

W Importer

Exporter 80 80

UK live F Importer

Exporter 25 55 80

W Importer 27 59 130 45 5 266

Exporter 27 83 170 90 73 443

skins W Importer

Exporter 26 3 29


Exporter 3 15 18

Malaysia Denmark live W Importer 5 25 30

Exporter 5 25 30

France live W Importer 5 15 30 50

Exporter 5 30 35

Germany live W Importer

Exporter 10 10


Exporter 7000 7000

Hungary live W Importer

Exporter 15 17 32

Spain live W Importer 25 75 26 45 171

Exporter 25 82 30 80 217


Exporter 5000 5000

Philippines France skins W Importer 99 99

Exporter

Singapore Germany specimens W Importer

Exporter

10 g (Purpose

S) 10

W Importer

Exporter

8 ml (Purpose

S) 8

W Importer

Exporter 4 (Purpose

S)

12 (Purpose

S) 16

United States UK small I Importer 14 14

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leather products

(Unspecified

purpose)

Exporter

Table 20. Main indirect exports of Varanus salvator to EU-27, 2001-2005.

Origin Term Source Purpose Reported by 2001 2002 2003 2004 2005 Total

Indonesia bodies U T Importer

Exporter 10 10

garments W T Importer 1 1

Exporter 2 2

large leather products C T Importer 4 4

Exporter 2 2

W T Importer 1042 34 8 90 49 1224

Exporter 61 7 37 10 8 123

live W Q Importer 2 2 4

Exporter 1 1

T Importer 6 5 11

Exporter 2 2 4

- Importer 100 100

Exporter 5 5

shoes W T Importer 755 76 831

Exporter 187 8 5 82 50 332

skin pieces C T Importer

Exporter 5 5

W T Importer


skins W P Importer

Exporter 4 4

T Importer 56454 17496 5489 67119 103967 250525

Exporter 52934 19077 8539 77976 108188 266714

- Importer

Exporter 3583 473 220 770 5046

small leather products F T Importer

Exporter 291 291

I - Importer 2300 2300

Exporter

O T Importer 25 25

Exporter

U T Importer 12 12

Exporter 2 2

W P Importer 5 5

Exporter 5 4 8 17

T Importer 41275 93056 65629 6847 5514 212321

Exporter 1766 237 1025 288 1463 4779

- Importer

Exporter 5820 3815 1473 1910 13018

- - Importer 4004 1910 1352 3100 10366

Exporter

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Origin Term Source Purpose Reported by 2001 2002 2003 2004 2005 Total

Malaysia large leather products W T Importer 133 10 70 213

Exporter 2 76 6 19 103

shoes W E Importer

Exporter 110 110

T Importer 72 2 74


skin pieces W T Importer

Exporter 10 10

skins W T Importer 36510 34513 8500 20621 81319 181463

Exporter 39010 29500 18500 76007 70324 233341

small leather products T Importer 52 52

Exporter 10 10

W P Importer

Exporter 4 2 5 11

T Importer 238 97 242 83 3756 4416

Exporter 140 40 35 134 25 374

- Importer

Exporter 206 179 193 45 623

- - Importer 600 600 1200

Exporter

Thailand small leather products W T Importer 1 10 310 321

Exporter 23 23

- Importer


United States

small leather products R T Importer 14 14

Exporter

Unknown skins I L Importer 1 1

Exporter 1 1 2

small leather products C T Importer 710 710

Exporter

I - Importer 187 2 189

Exporter

W T Importer 2250 2250

Exporter

- Importer

Exporter 115 115

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Table 21. Main direct exports of Varanus salvator to countries other than EU-27, 2001-2005.

Exporter Term Source Purpose Reported by Total

China small leather products W T Importer 335

Exporter

Hong Kong small leather products W T Importer 5034

Exporter

India skins W T Importer 1000

Exporter

Indonesia live C T Importer 7

Exporter 30

F T Importer 479

Exporter 1275

W P Importer 6

Exporter 2

Q Importer 14

Exporter

T Importer 16560

Exporter 22883

Z Importer 14

Exporter

meat W T Importer

Exporter 1000 kg

shoes W T Importer 1870

Exporter

skeletons W T Importer 59

Exporter

skin pieces W T Importer 3000 m, 9650 unspec. units

Exporter

skins A T Importer 400

Exporter

W T Importer 996623

Exporter 1926723

small leather products W P Importer 15

Exporter

Q Importer

Exporter 10

T Importer 15678

Exporter 22317

Malaysia gall W T Importer

Exporter 13

live C T Importer 6

Exporter 430

W B Importer 600

Exporter

P Importer 200

Exporter 200

Q Importer 84

Exporter 34

T Importer 55252, 20800 kg

Exporter 68232

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Exporter Term Source Purpose Reported by Total

Z Importer 36

Exporter 34

- Importer

Exporter 200

meat W T Importer 144104 kg

Exporter 154115 kg

shoes W T Importer 32

Exporter

skins C T Importer

Exporter 6000

W P Importer

Exporter 13000

T Importer 1024482

Exporter 916918

Philippines large leather products W T Importer 334

Exporter

skins W T Importer 13

Exporter

Singapore skins W T Importer 1100

Exporter

United States live C B Importer

Exporter 9

T Importer

Exporter 18

W Q Importer 4

Exporter

- - Importer 51

Exporter

small leather products C T Importer 14

Exporter

CONSERVATION STATUS IN RANGE STATES: V. salvator is a large, diurnal carnivorous monitor lizard, reaching up to 2.5m total length (Luxmoore et al., 1988). There are seven or more subspecies (Auliya, 2006). It occurs in the vicinity of water and also lives near human settlements (Luxmoore and Groombridge, 1990). Permanent surface water was found to be the most important factor determining its presence in Flores (Lesser Sundas) (Auffenberg, 1981) and apparently this applies throughout its range (Luxmoore et al., 1988). They have extensive home ranges and reproduction occurs biannually or throughout the year (Auliya, 2006). Clutch size increases with maternal body size (Shine et al., 1996) and is typically around 15 eggs, with up to 40 over the course of a year (Bennett, undated). Ranging from from South Asia to Eastern Indonesia (Auliya, 2006), V. salvator was said to be common or very common in the early twentieth century, and although reported to be depleted locally, it was still relatively common over much of its range (Luxmoore et al., 1988). Utilisation was reported to be a more important factor in the species' decline than habitat loss (Luxmoore et al., 1988). BANGLADESH: Occurrence reported in mangrove forest in the Sundarbans region, and other habitats near the Bay of Bengal, including Cox' Bazar and Chittagong areas and islands such as St Martin's (Luxmoore and Groombridge, 1990). Records suggest that it was formerly more

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widespread and has disappeared from everywhere but the coast (Luxmoore and Groombridge, 1990 and references therein). Protected by the Bangladesh Wild Life (Preservation) Order, 1973 (Order of the President No. 23 of 1973), which prohibits the export of skins and illegal trade has occurred in large volumes and is complicated by alleged cross-border movement of skins between India and Bangladesh (Luxmoore and Groombridge, 1990). Considered Endangered by IUCN Bangladesh (Khan et al., 2001). No exports were reported to CITES between 2001-2005. BRUNEI DARUSSALAM: Occurrence reported (Cranbrook & Edwards, 1994). Became Party to CITES in 1990. No exports were reported to CITES between 2001-2005. CAMBODIA: Became Party to CITES in 1997. No exports were reported to CITES between 2001-2005. CHINA: V. salvator salvator is restricted to southern China: Hainan; Guangdong (?); Yunnan; and Guangxi (Zhao and Adler, 1993). Given Level 1 (most restrictive) protection in the 1987 Guo Huan (National Environmental Protection Committee) number 014 “List of Important Protected Wild Animals”, reptiles and amphibian pages 11-12, amended 1989, pages 27-28. Observed in large quantities in major free trade markets and at frontier trade sites in Guangdong, Guangxi (Li et al., 1996) and Yunnan (Li and Wang, 1999). CITES-reported export trade of 335 small leather products between 2001-2005; no exports to EU-27 countries were reported. HONG KONG, CHINA: Occurrence reported by Romer (1963) but apparently no longer present here (CITES management authority, cited by Luxmoore and Groombridge, 1990). The 'Protection of Endangered Species Ordinance (Chapter 187, Ordinance no 63, 1976, revised 1987 & 1992) afforded total protection for several species, including V. salvator. Importer-reported trade of 5034 small leather products between 2001-2005 to countries other than EU-27; no exports to EU-27 countries were reported. INDIA: Occurs in Orissa, Bengal, eastern India (Daniel, 1983). The subspecies V. salvator salvator is found here, including the Nicobar Islands. V. salvator andamanensis is found in the Andaman Islands. Listed in Schedule I (completely protected species) of the Wildlife Preservation Act of 1972. In spite of this protection and the ban on trade in reptile skins in 1979, in 2001 large numbers of reptile products were still leaving the country (Das, 2001) Once widespread, common and abundant around the Andaman and Nicobar Islands, Varanus salvator andamanensis has become extremely rare and is under threat in the Andamans (Andrews, 2001) due to habitat loss and hunting for meat, fat and skin. Hunting is with dogs and noose traps and feral dogs are also a threat. Habitat destruction (especially mangrove clearance) and over-collecting for meat and skin seem to be responsible for the decline in numbers in mainland India (Bennett, undated). Import of 1000 skins in 2001 by Republic of Korea; no exports to EU-27 countries were reported. INDONESIA: Occurrence of V. salvator salvator reported in Sumatra, Nias, Engano, Banka, Borneo (Kalimantan) and Sulawesi; V. salvator bivittatus found in Java, Bali, Lombok, Sumbawa, Flores, Wetar and V. salvator togianus found in Timotto north of Sulawesi (Welch et al., 1990). Siak district (Anon, 2005). Distribution lists were compiled by Auliya (2006): 20th century records included Pulau Berhala (island East of N. Sumatra), Bawean Islands (N. of E. Java), Ringi Island and Sumba (Lesser Sundas). 19th century records included Timor, Seram and Timotoo Island (Togian Islands). Reports from researchers and reptile traders include Sulawesi and Selayar Island. Known to colonise remote islands, for example it was recorded on Krakatau six years after the 1883 eruption (Thornton, 1996). Auliya (2006) estimated a population density of 0.4-0.5 animals/ha on a 650 ha study site which he compared to recent (1979 or later) studies from other sites, including Philippines and

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Bangladesh which, according to Auliya (2006) found 0.9, 1.4 and 4.15-5.55 animals per 20ha (0.045, 0.07 and 0.206-0.2775/ha, respectively). Riquier (1998) gave preliminary population density estimates of 0.12/ha in West Kalimantan. Very large animals are not harvested for the leather industry due to the poor quality of their skins, which become thick and scarred (Shine et al., 1996). Shine et al. (1996) found that V. salvator persisted in southern Sumatra in spite of several decades of intensive harvesting, which they attributed to a large area of suitable habitat, ecological flexibility, high reproductive rate and perhaps commercial harvesting focussing on adult males. They considered that the levels of harvesting might cause extirpation from local areas but not cause extinction. Auliya (2006) commented that Varanus salvator is “definitely not vulnerable to becoming extinct” due to the extensive geographical distribution compared to endemic species. However, when comparing capture data from intensively harvested sites with non-harvested sites Auliya (2006) found higher class sizes (larger specimens) in non-harvested populations. Erdelen (1998) proposed measures which might be evaluated to lessen the impact of the skin trade: − the use of intermediate size classes only − no killing of accidentally caught specimens − uses other than skin to be limited to skinned specimens − no extension of export of live specimens for consumption − release of hatchlings from collected females into the location where caught − a shift from crust tanned skins to quality leather products LAO PEOPLE'S DEMOCRATIC REPUBLIC: A 1998 survey found they had become much rarer over the preceding 10 years in Phou Louey National Biodiversity Conservation Area, Houaphanh Province (Stuart, 1998). No exports were reported to CITES between 2001-2005. MALAYSIA: V. salvator is found in most habitats, although it is most common in and near rivers, lakes, swamps, mangroves and coastal wetlands, including small islands deficient in freshwater in the Malayan archipelago (Traeholt, 1998).

Grismer et al (2006) listed V. salvator as present in the following islands of the Tioman archipelago: Aceh, Anak Aceh, Layak, Tunas Selatan, Tunas Utara, Tengah, Pelandok, Mawar, Setindan, Batu Gajah, Seribuat, Sembilang, Tasu, Mertang Barat, Mertang Tengah, Rawa, Harimau, Mensirip, Babi Besar, Babi Hujong, Simbang, Sibu, Sibu Tengah, Sibu Hujong, Sibu Kukus, Tokong Belalai, Tioman, Tulai, Pemanggil, Aur, Dayang.

The 2005 annual report of the Department of Widlife and National Parks, Kuala Lumpur (DNWP, 2005) stated that the Water Monitor was the most hunted wild animal in 2005. It noted two court cases in 2005 involving:

1) illegal possession of Water Monitor and Clouded Monitor, which resulted in a 2,000 RM fine and 6 months imprisonment.

2) illegal possession of 3425 kg of meat of Water Monitor and 1140 kg of meat of python - fined 1,800 RM

Sabah: reportedly widespread and abundant wherever there is water (Luxmoore and Groombridge, 1990), low numbers were consumed by non-muslims but none were exported (Luxmoore and Groombridge, 1990). Monitor meat was uncommon in the markets (Traeholt, 1998). V. salvator gall bladders were used in Chinese medicine (Traeholt, 1998). Peninsular Malaysia: Jasmi (1988, cited by Luxmoore and Groombridge, 1990) reported that habitats remained widespread and populations were increasing, though Kiew (pers. comm.

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1989 to Luxmoore and Groombridge, 1990) doubted that numbers were increasing. Khan (1969, cited by Luxmoore and Groombridge, 1990) recorded exports from all eleven states of Peninsular Malaysia, suggesting distribution throughout the country. Sarawak: reported to be widespread and common, hunted intensively for food, but trade was small and insignificant (Luxmoore and Groombridge, 1990) Skins were exported in large numbers from Malaysia. Most trapping occurred in states with large non-muslim populations as Muslims are forbidden to trap lizards for skins commercially (Luxmoore and Groombridge, 1990). Included in Schedule II of the Protection of Wildlife Act No. 76 (1972) and Part I of Schedule 2 of the Wildlife Conservation Enactment (1997), which states that the species may not be hunted without a licence, for which the given quota must not be exceeded. Traeholt (1998) noted that 111,515 lizards were trapped under licence and 130,099 skins were exported. Traeholt recommends on-site checks and enforcement. Oil palm estates appeared to offer excellent habitat for the species, in which it can be found at higher densities than virgin forest (Traeholt, 1998). It was still abundant throughout most of its natural range in spite of substantial harvest by commercial trappers (Traeholt, 1998). It was likely to be affected here by habitat loss and cleaning of rivers for public health. The average body size in hunted areas was significantly smaller than non-hunted areas (Traeholt, 1998). MYANMAR: Under the Forest Department Notification No. 583/94 (1994), Varanus salvator is included in the list of species offered protection by the Protection of Wildlife, Wild Plants and Conservation of Natural Areas Act 15(A) (1994). No exports were reported to CITES between 2001-2005. PHILIPPINES: V. salvator cumingi was found on Mindanao, Leyte Cebu and Samar Mindanao, probably Bohol. V. salvator marmoratus found on Luzon, Culion, Calamianes, Palawan, probably Mindoro . V. salvator nuchalis found on Negros, Guimares, Cebu, Panay, Masbate (Welch et al., 1990; Luxmoore and Groombridge, 1990). Natural populations appeared to be 'strongly perturbed': not only were numbers low, but adults were very rare (Gaulke, 1988, cited by Luxmoore and Groombridge, 1990). This was attributed to habitat modification and loss coupled with intensive exploitation for food and the skin trade (Luxmoore and Groombridge, 1990). Listed in Annex A of DENR administrative order No. 48 (1991) Protection and Conservation of Philippine Wild Birds, Mammals, and Reptiles. This declares the species as a priority concern for conservation and protection and makes it unlawful for any person, groups or entity to collect and/or trade the species, provided that, such acts are covered by a permit granted by the DENR pursuant to the pertinent provisions of DENR AO 96, series of 1988 entitled,"Policies on the allocation of Quota on Wildlife". Thought to be the most heavily exploited reptile in the Philippines; the results of a questionnaire given to residents in Laguna, Quezon, Nueva Ecija and Bulcan provinces indicated that numbers are declining, probably associated with hunting pressure (Villamor, 1993). CITES-reported export trade of 334 large leather products and 13 skins between 2001-2005; no exports to EU-27 countries were reported to CITES. SINGAPORE: A small population, which was not known to be exploited (P. Gopalakrishnakone, in litt. cited by Luxmoore and Groombridge, 1990). Most of the world trade in V. salvator passes through Singapore, which acts as a clearing-house for skins from all over South East Asia.

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Importer-reported trade totalled 1100 skins between 2001-2005; no direct exports from Singapore to EU-27 countries were reported to CITES during this period. When considering indirect trade to EU-27 (Table 20), the main exporter was Singapore, accounting for a total of 423469 skins (source W, purpose P) as reported by importer, or 497042 skins (source W, purpose P) as reported by exporter. Slightly over half of those skins originated in Indonesia, and the rest in Malaysia. SRI LANKA: Occurrence of V. salvator salvator reported (Welch et al., 1990). Abundant and little persecuted (Luxmoore and Groombridge, 1990). Protected by the Flora and Fauna Protection Ordinance, Cap. 469 of 1938, amended Act 44 of 1964. No exports were reported to CITES between 2001-2005. THAILAND: Occurs throughout the country, associated with large bodies of water and other habitats wherever surface water is present. Available reports disparately describe the species as common or uncommon (Luxmoore and Groombridge, 1990). Hunting was reported to occur throughout the country, but was most concentrated in the south (Luxmoore and Groombridge, 1990). Trade reported to CITES comprised small leather products, indirectly imported to the EU totalling 321 reported by importer and 336 reported by exporter. VIET NAM: Listed as 'Vulnerable' in the Red Data Book for Viet Nam (Anon. (2000) cited by Truong and Bain. R, 2006). No exports were reported to CITES between 2001-2005.

REFERENCES: Andrews, H. (2001) Threatened Herpatofauna of Andaman and Nicobar Islands. In:

Bambaradeniya, C.N.B. and V.N. Samarasekara (Editors)(2001) An overview of the threatened herpatofauna of South Asia. IUCN Sri Lanka and Asia Regional Biodiversity Programme, Colombo, Sri Lanka vi+118 pp. Pages 39-47.

Anon (2005) Rainforest Alliance SmartWood Program High Conservation Value Forest (HCVF) Assessment Report for: Siak District. Final Report.

Auffenberg, W. (1981) The behavioral ecology of the Komodo monitor. University Presses of Florida, Gainesville, FL.

Auliya, M.A. (2006) Taxonomy, life history and conservation of giant reptiles in West Kalimantan. Verlag GmbH, Münster.

Bennett, D. undated. The water monitor, Varanus salvator. First Published in Reptilian 3(8).from URL:http://www.cyclura.com/modules.php?op=modload&name=News&file=article&sid=104&mode=thread&order=0&thold=0 Accessed July 2007.

Cranbrook, Earl of and Edwards, D. S. (1994) Belalong: a tropical rainforest. The Royal Geographical Society, London & Sun Tree Publishing, Singapore. 389 pp.

Daniel, J.C. (1983) The book of Indian reptiles. Bombay Natural History Society and Oxford University Press, Oxford. 141 pp.

Das, I. (2001) Threatened herpatofauna of India. In: Bambaradeniya, C.N.B. and V.N. Samarasekara (Editors)(2001) An overview of the threatened herpatofauna of South Asia. IUCN Sri Lanka and Asia Regional Biodiversity Programme, Colombo, Sri Lanka vi+118 pp. Pages 63-70.

DNWP (2005) Annual report. Department of Widlife and National Parks, Peninsular Malaysia. Kuala Lumpur. Malaysia. http://www.wildlife.gov.my/ accessed 07/08/2007.

Erdelen, W. (1998) Trade in lizards and snakes in Indonesia: bio-geography, ignorance, and sustainability. Mertensiella 9: 69-83.

Grismer, L.L., Youmans, T.M. Wood, P.L. and Grismer, J.L. (2006) Checklist of the herpetofauna of the seribuat archipelago, west malaysia with comments on biogeography, natural history, and adaptive types. The Raffles Bulletin of Zoology, 54(1): 157-180.

Khan, M.A., Khan, M.H. and Chowdhury, M.M. (2001) Threatened herpatofauna of Bangladesh In: Bambaradeniya, C.N.B. and V.N. Samarasekara (Editors)(2001) An

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overview of the threatened herpatofauna of South Asia. IUCN Sri Lanka and Asia Regional Biodiversity Programme, Colombo, Sri Lanka vi+118 pp. Pages 48-58.

Li, W., Fuller, T.K. and Wang, S. (1996). A survey of wildlife trade in Guangxi and Guangdong, China. TRAFFIC Bulletin 16(1):9-16.

Li, W and Wang, H. (1999) Wildlife Trade in Yunnan Province, China, at the Border with Viet Nam. TRAFFIC Bulletin 18(1):22-31

Luxmoore, R. and Groombridge, B. (1990) Asian Monitor Lizards: a review of distribution, status, exploitation and trade in four selected species. Secretariat of the Convention on International Trade in Endangered Species of Wild Fauna and Flora, Lausanne, Switzerland. 195 pp.

Luxmoore, R., Groombridge, B. and Broad, S. (1988) The significance of trade in selected species listed in CITES Appendix II. Vol. 3. Reptiles and invertebrates. Lausanne: CITES Secretariat.

Riquier, M.A. (1998) Status, population biology and conservation of the water monitor (Varanus salvator), the reticulated python (Python reticulatus), and the blood python (Python curtus) in Sumatra and Kalimantan, Indonesia – project report Kalimantan. Mertensiella 9:119-129.

Romer, J. D. (1963) Occurrence of the Common Water Monitor Varanus salvator (Laurenti), in Hong Kong. pp. 12. Memoirs of the Hong Kong Natural History Society.

Shine, R., Harlow, P.S. and Keogh, J.S. (1996) Commercial harvesting of giant lizards: the biology of water monitors Varanus salvator in southern Sumatra. Biological Conservation 77:125-134.

Stuart, B. (1998) A Survey of Amphibians and Reptiles in Phou Louey National Biodiversity Conservation Area, Houaphanh Province, Lao PDR. Wildlife Conservation Society Lao Program, Vientiane, Lao PDR.

Thornton, I.W.B. (1996) Krakatau: the destruction and reassembly of an island ecosystem Harvard University Press, Cambridge, Mass. 346 pp.

Traeholt, C. (1998) Exploitation and trade of the water monitor lizard (Varanus salvator) in Malaysia. Mertensiella 9: 131-135.

Truong, N.Q. and Bain, R. (2006). An Assessment of the Herpetofauna of the Green Corridor Forest Landscape, Thua Thien Hue Province, Viet Nam. Report No 2: Green Corridor Project, WWF Greater Mekong & Viet Nam Country Programme and FPD Thua Thien Hue Province, Viet Nam.

UNEP-WCMC 2006. Analysis of the European Community and Candidate Countries Annual Reports to CITES 2004. Prepared for the European Commission, Directorate General E- Environment ENV E.4- Development and Environment pp 163.

Villamor, C.I. (1993) Morphometry and conservation status of water monitor lizard (Varanus salvator) in the Philippines. Asia Life Sciences 2(2):113-120.

Welch, K.R.G., Cooke, P.S. and Wright, A.S. (1990) Lizards of the Orient: a checklist. Robert E. Krieger publishing company. Malabar, Florida. 162 pp.

Zhao, E.M. and Adler, K. (1993) Herpetology of China. Society for the study of Amphibians and Reptiles.

FURTHER READING: Alcala, A. C. 1986. Guide to Philippine flora and fauna: amphibians and reptiles. Quezon

City: Natural Resources Management Center, Ministry of Natural Resources and University of the Philippines.

Biswas, S. and Kar, S. 1981. Some observations on nesting habitats and biology of Varanus salvator (Laurenti) of Bhitarkanika Sanctuary, Orissa. pp. 303-308. Journal of the Bombay Natural History Society.

Böhme, W. 2003. Checklist of the living monitor lizards of the world (family Varanidae). pp. 3-43.. Zoologische Verhandelingen.

Cox, C. R., Vu Van Dung and Pham Mong Giao. 1992. Report of a management feasability study of the Muong Nhe Nature Researve (November/December 1991).: 75.

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Cranbrook, Earl of and Edwards, D. S. 1994. A tropical rainforest. The nature of biodiversity in Borneo at Belalong, Brunei. London & Singapore: The Royal Geographical Society & Sun Tree Publishing.

Daniel, J. C. 1983. The book of Indian reptiles. Khan, M. A. R. 1982. Wildlife of Bangladesh. Dhaka: University of Dhaka. Lekagul, B. 1969. Monitors (_Varanus_) of Thailand. pp. 31-32.. Conservation News, S.E. Asia. Luxmoore, R. and Groombridge, B. 1990. Asian Monitor lizards: a review of distribution,

status, exploitation and trade in four selected species. Lausanne: CITES Secretariat. Luxmoore, R., Groombridge, B. and Broad, S. 1988. The significance of trade in selected

species listed in CITES Appendix II. Vol. 3. Reptiles and invertebrates. Lausanne: CITES Secretariat.

Mertens, R. 1942. Die Familie der Warane (Varanidae). Dritter Teil: Taxonomie. pp. 235-391.. Abhandlungen der Senckenbergischen Naturforschenden Gesellschaft.

Ng, P. K. L. & Wee, Y. C (eds.). 1994. The Singapore Red Data Book. Threatened Plants and Animals of Singapore. Singapore: The Nature Society (Singapore).

Pope, C. H. 1935. The reptiles of China. Natural history of Central Asia, X. American Museum of Natural History.

Ridley, H. N. 1899. The habits of Malay reptiles. pp. 185-210.. Journal of the Straits Branch of the Royal Asiatic Society.

Romer, J. D. 1963. Occurrence of the Common Water Monitor _Varanus salvator_ (Laurenti), in Hong Kong. pp. 12.. Memoirs of the Hong Kong Natural History Society.

Rooij, N. de. 1915. The reptiles of the Indo-Australian Archipelago. Vol. I. Lacertidae, Chelonia, Emydosauria. Leiden: E.J. Brill Ltd.

Smith, M. A. 1943. The fauna of British India. Reptilia and Amphibia, Vol. III, Serpentes. London : Taylor and Francis.

Stuart, B. 2001. Amphibians and reptiles of Cambodia. Hanoi: WWF Indochina Program. Stuart, B. L. 1999. Amphibians and reptiles. Vientiane, Lao PDR: IUCN. 43-67. Taylor, E. H. 1963. The lizards of Thailand. pp. 687-1077. University of Kansas Science

Bulletin.Traeholt, C. 1998. Exploitation and trade of the water monitor lizard (Varanus salvator) in Malaysia. Mertensiella 9: 131-135

Zug, G., Slowinski, J. and Wogan, G. 2003. Checklist of the amphibians and reptiles of Myanmar.

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EUPHORBIACEAE

SPECIES: Euphorbia antisyphilitica

SYNONYMS: Euphorbia cerifera, Tirucallia antisyphilitica,

Trichosterigma antisyphiliticum COMMON NAMES: candelilla (English) RANGE STATES: Mexico, United States. IUCN RED LIST: Not Evaluated PREVIOUS EC OPINIONS: -

TRADE: Euphorbia antisyphilitica was selected for this review following an overall increase and high volume of trade in wild sourced wax into the European Community over the period 1995-2004 (UNEP-WCMC, 2006).

No CITES export quotas have been published for Euphorbia antisyphilitica. Table 22. Main direct exports of Euphorbia antisyphilitica wax (in kg) to EU-27, 2001-2005. Source W, Term wax.

Exporter Importer Reported by 2001 2002 2003 2004 2005 Total

Mexico Germany Importer 41000 195000 191525 218500 231525 877550

Exporter 98000 208000 175000 232725 224000 937725

Table 23. Indirect exports of Euphorbia antisyphilitica to EU-27, 2001-2005. Source W. Exporters were Japan and Switzerland.

Origin Term Reported by 2004 2005 Total

Mexico Extract (kg) Importer 147 224 371

Exporter

Wax (kg) Importer

Exporter 273 273

Table 24. Direct exports of Euphorbia antisyphilitica to countries other than EU-27, 2001-2005. Source W unless otherwise indicated

Exporter Term Reported by Total

Mexico live Importer 5 (source I)

Exporter

Wax (kg) Importer 400

Exporter 383426

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Prior to 2001 the only reported trade was in live specimens. Exports from Mexico were: 4 in 1990; exports from the USA: 1 in 1983, 4 in 1988, 12 in 1990, 1 in 1991. Since the main trade is reported only as kilos of wax it is very difficult to relate the quantities to the number of plants involved. Using an estimate of wax yield of 2% of plant weight (see below) the 2005 reported exports of 224,000 kg would represent about 11 m kg plants and, using a maximum estimate of 25,000 kg of plants per hectare (derived from figures given below) leads to a minimum figure of 440 ha of habitat that were cleared to produce this amount of wax. However, if the annual production figure given by Bartsch (2004) is applicable to 2005 it could mean that nearly 6,000 ha of habitat were cleared. There are apparently no estimates of the areas of habitat occupied by this species so it is not possible to estimate the effects of trade on this scale.

CONSERVATION STATUS IN RANGE STATES: MEXICO: The following information was taken from Bartsch (2004): ‘Candelilla Euphorbia antisyphilitica occurs in Mexico and southern USA, with its principal habitat in the Chihuahuan Desert, 80 per cent of which lies within Mexico. The species is collected from the wild in Mexico for the production of a high quality wax known as candelilla wax. Like all succulent representatives of the genus, Euphorbia antisyphilitica is listed in CITES Appendix II owing to the demand for some Euphorbia succulent species in the live ornamental plant trade. The most intensive harvesting of Euphorbia antisyphilitica takes place in the Mexican Federal State of Coahuila. The wax is processed for application in many products, including cosmetics, food additives, and as a separation wax and/or mould during the production of certain goods. Apart from a very limited replanting of the roots of harvested specimens, there is no cultivation of the species for wax production. Between 60-80% of the annual wax production in Mexico is exported; the remainder is for domestic consumption. According to FAO data and information from traders, annual production amounts to approximately 3000 t a year. However, CITES annual report data show far less wax is being exported from Mexico (98 t of “extract” - a Customs declaration code used to describe the wax) in 2001 and 208 t of “wax” in 2002). Principal importers are the USA (1000 t/y) and the European Union (350 t/y). German imports of 175.5 t (2003) and 211 t (2002) are mostly for re-export, with a relatively small annual national demand of approximately 50 t/y, according to traders in 2004 (Bartsch, 2004). Harvesting regulations have been developed in Mexico, but are not applied consistently in rural areas. It is unclear if present levels of exploitation are a significant threat to Euphorbia antisyphilitica at the species level at this time; although there are no population estimates available, the species is still considered common in many parts of its range. However, owing to the direct and negative effect of use without management, many other populations of this species have declined or disappeared on a local and regional scale. Therefore, if present levels of harvesting and export continue without harvesting regulations being consistently applied and enforced by the Mexican authorities, this situation is likely to become a matter for significant concern. Despite the CITES listing of Euphorbia antisyphilitica, candelilla wax imports are not recorded by most countries, with the exception of Mexico and Germany. It is possible that the connection between “Candelilla wax” and a Euphorbia species has not been made by other countries and, therefore, the CITES relevance has been overlooked by CITES enforcement authorities. Until sustainable sourcing of Euphorbia antisyphilitica in its Mexican range is ensured, a downlisting [i.e. deletion] of the species from Appendix II, as has been suggested by several businesses, is not recommended. Furthermore, Mexican botanists regard the CITES listing as an important instrument for controlling harvesting and protection of the species in the wild.’ The following extract from the Candelilla Institute website suggests that some steps are being taken to manage the trade in this species but it is not clear what has been achieved so far:

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‘Sustainable Development Projects

Exploitation of the Candelilla plant has been going on for more than a century without an integral plan for utilization of this resource. The precarious living conditions of the inhabitants of the region, as well as the lack of other economic activities which could contribute to improving their income levels, have caused the indiscriminate exploitation of the Candelilla plant. This brings with it the threat of the gradual extinction of this important natural resource. On the other hand, the competition that prevails in the Candelilla wax industry has motivated companies to carry out investments in productive infrastructure in order to assure the supply of this important natural raw material. However, none of these investments have focused on the integral utilization of this resource through technological development and sustainable development programs, which could be efficiently put into practice. The problem, which has existed for over a century, remains unsolved. Sustainable development means making nature the base of the productive chain, not only as a set of ecological norms which the economy must assimilate, but above all as the driving force which promotes the development of all the productive forces arising from the ecological, cultural and geographic conditions of each region. Candelilla Institute is working toward this goal, supported by its technological associates and government agencies, which are committed to the economic and social development of the Candelilla Region. The following projects for sustainable development have been defined as an essential part of the integral program being carried out by Candelilla Institute:

• Analysis of the Candelilla plant inventory, cultivation programs and efficient exploitation techniques, which will guarantee the survival and conservation of the Candelilla plant.

• Implementation of economic and social programs, which will permit an improved lifestyle and living conditions of the inhabitants of the Candelilla Region.

The complexity of the problem must not be ignored. To achieve significant change in an industry that has remained unaltered for more than a century requires great effort and dedication. These are just the first steps toward creating a natural, technological and human competitive advantage in the Candelilla wax industry.’ (Candelilla Institute, 2004) USA: Robbins (2003) reported that the species was ‘Critically imperiled’ in New Mexico, that its status in Texas was unknown, but that globally it was ‘Widespread, abundant, secure’.

REFERENCES: Anon. 2004. Wax camps: from desert plants to dollars: candelilla, wax making, and wax

products. URL: http://www.texasbeyondhistory.net/waxcamps/techniques.html Downloaded 30 July 2007.

Bartsch, F. 2004. Preliminary assessment of the trade and use of Euphorbia antisyphilitica. Traffic Bulletin 20(1): 6.

Candelilla Institute. 2004. Candelilla. URL: http://www.candelilla.org/candelilla.htm Downloaded 30 July 2007.

Robbins, C. S. 2003. Chihuahuan desert cacti in the United States: an assessment of trade, management, and conservation priorities. Part I in C. S. Robbins, ed. Prickly trade: trade and conservation of Chihuahuan desert cacti. Washington DC: TRAFFIC North America.

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