74954 Federal Register / Vol. 79, No. 241 / Tuesday, December 16, 2014 / Proposed Rules 

DEPARTMENT OF COMMERCE 

National Oceanic and Atmospheric 

Administration 

50 CFR Parts 223 and 224 

[Docket No. 140707555–4999–01] 

RIN 0648–XD370 

Endangered and Threatened Wildlife 

and Plants; 12-Month Finding for the 

Eastern Taiwan Strait Indo-Pacific 

Humpback Dolphin, Dusky Sea Snake, 

Banggai Cardinalfish, Harrisson’s 

Dogfish, and Three Corals Under the 

Endangered Species Act 

AGENCY: National Marine Fisheries 

Service (NMFS), National Oceanic and 

Atmospheric Administration (NOAA), 

Commerce. 

ACTION: Proposed rule; 12-month 

petition finding; request for comments. 

SUMMARY: We, NMFS, have completed 

comprehensive status reviews under the 

Endangered Species Act (ESA) for seven 

foreign marine species in response to a 

petition to list those species. These 

seven species are the Eastern Taiwan 

Strait population of Indo-Pacific 

humpback dolphin (Sousa chinensis), 

dusky sea snake (Aipysurus fuscus), 

Banggai cardinalfish (Pterapogon 

kauderni), Harrisson’s dogfish 

(Centrophorus harrissoni), and the 

corals Cantharellus noumeae, 

Siderastrea glynni, and Tubastraea 

floreana. We have determined that the 

Eastern Taiwan Strait Indo-Pacific 

humpback dolphin is not a distinct 

population segment and therefore does 

not warrant listing. We have determined 

that, based on the best scientific and 

commercial data available, and after 

taking into account efforts being made 

to protect the species, Pterapogon 

kauderni, and Centrophorus harrissoni 

meet the definition of a threatened 

species; and Aipysurus fuscus, 

Cantharellus noumeae, Siderastrea 

glynni, and Tubastraea floreana meet 

the definition of an endangered species. 

Therefore, we propose to list these six 

species under the ESA. We are not 

proposing to designate critical habitat 

for any of the species proposed for 

listing, because the geographical areas 

occupied by these species are entirely 

outside U.S. jurisdiction, and we have 

not identified any unoccupied areas that 

are currently essential to the 

conservation of any of these species. We 

are soliciting comments on our 

proposals to list the six species. We are 

also proposing related administrative 

changes to our lists of threatened and 

endangered species. 

DATES: Comments on our proposed rule 

to list eight species must be received by 

February 17, 2015. Public hearing 

requests must be made by January 30, 

2015. 

ADDRESSES: You may submit comments 

on this document, identified by NOAA– 

NMFS–2014–0083, by any of the 

following methods: 

• Electronic Submissions: Submit all 

electronic public comments via the 

Federal eRulemaking Portal. Go to 

www.regulations.gov/

#!docketDetail;D=NOAA-NMFS-2014- 

0083. Click the ‘‘Comment Now’’ icon, 

complete the required fields, and enter 

or attach your comments. 

• Mail: Submit written comments to, 

Lisa Manning, NMFS Office of Protected 

Resources (F/PR3), 1315 East West 

Highway, Silver Spring, MD 20910, 

USA. 

Instructions: You must submit 

comments by one of the above methods 

to ensure that we receive, document, 

and consider them. Comments sent by 

any other method, to any other address 

or individual, or received after the end 

of the comment period, may not be 

considered. All comments received are 

a part of the public record and will 

generally be posted for public viewing 

on http://www.regulations.gov without 

change. All personal identifying 

information (e.g., name, address, etc.), 

confidential business information, or 

otherwise sensitive information 

submitted voluntarily by the sender will 

be publicly accessible. We will accept 

anonymous comments (enter ‘‘N/A’’ in 

the required fields if you wish to remain 

anonymous). Attachments to electronic 

comments will be accepted in Microsoft 

Word, Excel, or Adobe PDF file formats 

only. 

You can obtain the petition, status 

review reports, the proposed rule, and 

the list of references electronically on 

our NMFS Web site at http://

www.nmfs.noaa.gov/pr/species/

petition81.htm. 

FOR FURTHER INFORMATION CONTACT: Lisa 

Manning, NMFS, Office of Protected 

Resources (OPR), (301) 427–8403. 

SUPPLEMENTARY INFORMATION: 

Background 

On July 15, 2013, we received a 

petition from WildEarth Guardians to 

list 81 marine species as threatened or 

endangered under the Endangered 

Species Act (ESA). This petition 

included species from many different 

taxonomic groups, and we prepared our 

90-day findings in batches by taxonomic 

group. We found that the petitioned 

actions may be warranted for 27 of the 

81 species and announced the initiation 

of status reviews for each of the 27 

species (78 FR 63941, October 25, 2013; 

78 FR 66675, November 6, 2013; 78 FR 

69376, November 19, 2013; 79 FR 9880, 

February 21, 2014; and 79 FR 10104, 

February 24, 2014). This document 

addresses the findings for 7 of those 27 

species: the Eastern Taiwan Strait 

population of Indo-Pacific humpback 

dolphin (Sousa chinensis), dusky sea 

snake (Aipysurus fuscus), Banggai 

cardinalfish (Pterapogon kauderni), 

Harrisson’s dogfish (Centrophorus 

harrissoni), and the corals Cantharellus 

noumeae, Siderastrea glynni, and 

Tubastraea floreana. The remaining 20 

species will be addressed in subsequent 

findings. 

We are responsible for determining 

whether species are threatened or 

endangered under the ESA (16 U.S.C. 

1531 et seq.). To make this 

determination, we consider first 

whether a group of organisms 

constitutes a ‘‘species’’ under the ESA, 

then whether the status of the species 

qualifies it for listing as either 

threatened or endangered. Section 3 of 

the ESA defines a ‘‘species’’ to include 

‘‘any subspecies of fish or wildlife or 

plants, and any distinct population 

segment of any species of vertebrate fish 

or wildlife which interbreeds when 

mature.’’ On February 7, 1996, NMFS 

and the U.S. Fish and Wildlife Service 

(USFWS; together, the Services) adopted 

a policy describing what constitutes a 

distinct population segment (DPS) of a 

taxonomic species (the DPS Policy; 61 

FR 4722). The DPS Policy identified two 

elements that must be considered when 

identifying a DPS: (1) The discreteness 

of the population segment in relation to 

the remainder of the species (or 

subspecies) to which it belongs; and 

(2) the significance of the population 

segment to the remainder of the species 

(or subspecies) to which it belongs. As 

stated in the DPS Policy, Congress 

expressed its expectation that the 

Services would exercise authority with 

regard to DPSs sparingly and only when 

the biological evidence indicates such 

action is warranted. 

Section 3 of the ESA defines an 

endangered species as ‘‘any species 

which is in danger of extinction 

throughout all or a significant portion of 

its range’’ and a threatened species as 

one ‘‘which is likely to become an 

endangered species within the 

foreseeable future throughout all or a 

significant portion of its range.’’ We 

interpret an ‘‘endangered species’’ to be 

one that is presently in danger of 

extinction. A ‘‘threatened species,’’ on 

the other hand, is not presently in 

danger of extinction, but is likely to 

become so in the foreseeable future (that 

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is, at a later time). In other words, the 

primary statutory difference between a 

threatened and endangered species is 

the timing of when a species may be in 

danger of extinction, either presently 

(endangered) or in the foreseeable future 

(threatened). 

When we consider whether species 

might qualify as threatened under the 

ESA, we must consider the meaning of 

the term ‘‘foreseeable future.’’ It is 

appropriate to interpret ‘‘foreseeable 

future’’ as the horizon over which 

predictions about the conservation 

status of the species can be reasonably 

relied upon. The foreseeable future 

considers the life history of the species, 

habitat characteristics, availability of 

data, particular threats, ability to predict 

threats, and the reliability to forecast the 

effects of these threats and future events 

on the status of the species under 

consideration. Because a species may be 

susceptible to a variety of threats for 

which different data are available, or 

which operate across different time 

scales, the foreseeable future is not 

necessarily reducible to a particular 

number of years. Discussions of the 

considerations for each relevant species 

are in the species-specific sections 

below. 

Section 4(a)(1) of the ESA requires us 

to determine whether any species is 

endangered or threatened due to any 

one or a combination of the following 

five threat factors: The present or 

threatened destruction, modification, or 

curtailment of its habitat or range; 

overutilization for commercial, 

recreational, scientific, or educational 

purposes; disease or predation; the 

inadequacy of existing regulatory 

mechanisms; or other natural or 

manmade factors affecting its continued 

existence. We are also required to make 

listing determinations based solely on 

the best scientific and commercial data 

available, after conducting a review of 

the species’ status and after taking into 

account efforts being made by any state 

or foreign nation to protect the species. 

In making a listing determination, we 

first determine whether a petitioned 

species meets the ESA definition of a 

‘‘species.’’ Next, using the best available 

information gathered during the status 

review for the species, we complete a 

status and extinction risk assessment. In 

assessing extinction risk, we consider 

the demographic viability factors 

developed by McElhany et al. (2000) 

and the risk matrix approach developed 

by Wainwright and Kope (1999) to 

organize and summarize extinction risk 

considerations. The approach of 

considering demographic risk factors to 

help frame the consideration of 

extinction risk has been used in many 

of our status reviews, including for 

Pacific salmonids, Pacific hake, walleye 

pollock, Pacific cod, Puget Sound 

rockfishes, Pacific herring, scalloped 

hammerhead sharks, and black abalone 

(see http://www.nmfs.noaa.gov/pr/

species/ for links to these reviews). In 

this approach, the collective condition 

of individual populations is considered 

at the species level according to four 

demographic viability factors: 

Abundance, growth rate/productivity, 

spatial structure/connectivity, and 

diversity. These viability factors reflect 

concepts that are well-founded in 

conservation biology and that 

individually and collectively provide 

strong indicators of extinction risk. 

We then assess efforts being made to 

protect the species, to determine if these 

conservation efforts are adequate to 

mitigate the existing threats. Section 

4(b)(1)(A) of the ESA requires the 

Secretary, when making a listing 

determination for a species, to take into 

consideration those efforts, if any, being 

made by any State or foreign nation to 

protect the species. We also evaluate 

conservation efforts that have not yet 

been fully implemented or shown to be 

effective using the criteria outlined in 

the joint NMFS/USFWS Policy for 

Evaluating Conservation Efforts (PECE; 

68 FR 15100, March 28, 2003), to 

determine their certainty of 

implementation and effectiveness. The 

PECE is designed to ensure consistent 

and adequate evaluation of whether any 

conservation efforts that have been 

recently adopted or implemented, but 

not yet demonstrated to be effective, 

will result in recovering the species to 

the point at which listing is not 

warranted or contribute to forming the 

basis for listing a species as threatened 

rather than endangered. The two basic 

criteria established by the PECE are: (1) 

The certainty that the conservation 

efforts will be implemented; and (2) the 

certainty that the efforts will be 

effective. We consider these criteria in 

each species-specific section, as 

applicable, below. Finally, we re-assess 

the extinction risk of the species in light 

of the existing conservation efforts. 

Status Reviews 

Status reviews for the petitioned 

species addressed in this finding were 

conducted by NMFS OPR staff. Separate 

status reviews were done for the Eastern 

Taiwan Strait Indo-Pacific humpback 

dolphin (Whittaker, 2014), dusky sea 

snake (Manning, 2014), Banggai 

cardinalfish (Conant, 2014), Harrison’s 

dogfish (Miller, 2014), and the three 

corals (Meadows, 2014). In order to 

complete the status reviews, we 

compiled information on the species’ 

biology, ecology, life history, threats, 

and conservation status from 

information contained in the petition, 

our files, a comprehensive literature 

search, and consultation with experts. 

We also considered information 

submitted by the public in response to 

our petition findings. Draft status review 

reports were also submitted to 

independent peer reviewers; comments 

and information received from peer 

reviewers were addressed and 

incorporated as appropriate before 

finalizing the draft reports. 

Each status review report provides a 

thorough discussion of demographic 

risks and threats to the particular 

species. We considered all identified 

threats, both individually and 

cumulatively, to determine whether the 

species responds in a way that causes 

actual impacts at the species level. The 

collective condition of individual 

populations was also considered at the 

species level, according to the four 

demographic viability factors discussed 

above. 

The status review reports are available 

on our Web site (see ADDRESSES

section). Below we summarize 

information from those reports and the 

status of each species. 

Eastern Taiwan Strait Population of the 

Indo-Pacific Humpback Dolphin 

The following section describes our 

analysis of the status of the Eastern 

Taiwan Strait (ETS) population of the 

Indo-Pacific Humpback dolphin, Sousa 

chinensis. 

Species Description 

The Indo-Pacific humpback dolphin, 

Sousa chinensis (Osbeck, 1765), within 

the genus Sousa, family Delphinidae, 

and order Cetacea, is broadly 

distributed. The taxonomy of the genus 

is unresolved and has historically been 

based on morphology, but genetic 

analyses have recently been used. 

Current taxonomic hypotheses identify 

Sousa chinensis as one of two (Jefferson 

et al., 2001), three (Rice, 1998), or four 

(Mendez et al., 2013) species within the 

genus. Each species is associated with a 

unique geographic range, though the 

species’ defined ranges vary depending 

on how many species are recognized. 

Rice (1998) recognizes Sousa teuzii in 

the eastern Atlantic, Sousa plumbea in 

the western Indo-Pacific, and Sousa 

chinensis in the eastern Indo-Pacific. 

Mendez et al. (2013) recently identified 

an as-yet unnamed potential new 

species in waters off of northern 

Australia. Currently, the International 

Union for Conservation of Nature 

(IUCN) and International Whaling 

Commission (IWC) Scientific Committee 

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recognize only two species, Sousa 

chinensis in the Indo-Pacific, and Sousa 

teuzii in the eastern Atlantic. Here, we 

follow a similar two-species taxonomy 

in our consideration of the genus and 

identification of the species Sousa 

chinensis. Under that taxonomy, Sousa 

chinensis’ range includes nearshore 

tropical and subtropical habitats in 

southern Africa, the Indian Ocean, 

North Australia, southern mainland 

China, Hong Kong, and Taiwan 

(Jefferson et al., 2001; Mendez et al., 

2013). We chose to follow a two-species 

taxonomy as it provides the clearest 

genetic, morphological, and geographic 

delineation of the species and is well 

supported by the current data available. 

While growing genetic and 

phylogeographic evidence suggests that 

Sousa chinensis is associated with 

further genetic subdivisions, more data 

are needed to clarify the taxonomy and 

delineate the geographic boundaries and 

ranges of these additional genetic units 

(Cockroft et al., 1997; Jefferson et al., 

2004b; Fre`re et al., 2008; Fre`re et al., 

2011; Lin et al., 2012; Mendez et al., 

2013). 

The Indo-Pacific humpback dolphin 

is easy to distinguish from other 

dolphin species in its range, as it is 

characterized by a robust body, a long, 

distinct beak, a short dorsal fin atop a 

wide dorsal hump, and round-tipped, 

broad flippers and flukes (Jefferson et 

al., 2001). The Indo-Pacific humpback 

dolphin is medium-sized, up to 2.8 m in 

length, weighing 250–280 kg (Ross et 

al., 1994). Morphological plasticity 

exists among populations of the species 

and is correlated with their geographic 

distributions (Ross et al., 1994). For 

example, the Eastern Taiwan Strait 

population, which occurs at the eastern 

portion of the species’ range, has a short 

dorsal fin with a wide base; the base of 

the fin measures 5–10 percent of the 

body length and slopes gradually into 

the surface of the body. This differs 

from individuals in the western portion 

of the range, which have a larger hump 

that comprises about 30 percent of body 

width, and forms the base of an even 

smaller dorsal fin (Ross et al., 1994). 

Males and females from the Pearl River 

Estuary population, and in other 

populations of Southeast Asia, do not 

exhibit sexual dimorphism in size, 

growth patterns, or morphology 

(Jefferson et al., 2001; Jefferson et al., 

2012). In contrast, individuals from 

South Africa exhibit sexual dimorphism 

in terms of size and dorsal hump 

morphology (Ross et al., 1994; 

Karczmarski et al., 1997). 

The species occurs in a range of 

nearshore habitats, including estuaries, 

mangroves, seagrass meadows, coastal 

lagoons, and sandy beaches (Ross et al., 

1994). In Thailand, Malaysia, and 

Indonesia, nearshore ecosystems are 

associated with tropical seagrass, coral, 

and mangrove lagoons (Beasley et al., 

1997; Smith et al., 2003; Adulyanukosol 

et al., 2006; Jaroensutasinee et al., 2011; 

Cherdsukai et al., 2013). In India, the 

species is associated with nearshore 

habitat consisting of mangroves, corals, 

and tidal mudflat, heavily influenced by 

monsoons that regulate the influx of 

freshwater to the system (Sutaria et al., 

2004). The coast of mainland China is 

thought to host at least eight 

populations of the species, primarily 

occurring in estuarine systems at the 

mouths of large rivers (Jefferson et al., 

2001; Jefferson et al., 2004a). Two 

coastal Chinese populations, in close 

proximity to the population in the 

Eastern Taiwan Strait, are relatively 

well-studied. These are the Pearl River 

Estuary/Hong Kong population and the 

Jiulong River Estuary/Xaimen 

population, both of which depend upon 

ecosystem productivity associated with 

the nutrient output supplied by large 

rivers (Chen et al., 2008; Chen et al., 

2010). 

The Eastern Taiwan Strait population 

of Sousa chinensis (henceforth referred 

to as the ETS humpback dolphin), for 

which we were petitioned, was first 

described in 2002 during an exploratory 

survey of coastal waters off of western 

Taiwan (Wang et al., 2004). Prior to 

these coastal surveys, there are few 

records mentioning the species in this 

region, save two strandings, a few 

photographs, and anecdotal reports 

(Wang, 2004), so their history in the 

region is unclear. Since the first survey 

in 2002, researchers have confirmed 

their year-round presence in the Eastern 

Taiwan Strait (Wang et al., 2011), 

inhabiting estuarine and coastal waters 

of central-western Taiwan. 

The ETS humpback dolphin habitat is 

most similar to that of the populations 

located off the coast of mainland China. 

Individuals of the ETS humpback 

dolphin population are thought to be 

restricted to water less than 30 meters 

deep, and most observed sightings have 

occurred in estuarine habitat with 

significant freshwater input (Wang et 

al., 2007b). Across the ETS humpback 

dolphin habitat, bottom substrate 

consists of soft-sloping muddy sediment 

with elevated nutrient inputs, primarily 

influenced by river deposition (Sheehy, 

2010). These nutrient inputs support 

high primary production, which fuels 

upper trophic levels, contributing to the 

dolphin’s source of food (Jefferson, 

2000). 

The Indo-Pacific humpback dolphin 

is considered a generalist and 

opportunistic piscivore (Barros et al., 

2004). As is common to the species as 

a whole, the ETS population uses 

echolocation and passive listening to 

find its prey. While little is known 

about the specific diet and feeding of 

the ETS population, diet can be inferred 

from that of other humpback dolphin 

populations (Barros et al., 2004; Chen et 

al., 2009). In Chinese waters off Hong 

Kong, the species consumes both 

bottom-dwelling and pelagic fish 

species, including croakers (Sciaenidae), 

mullets (Mugilidae), threadfins 

(Polynemidae), and herring (Clupeidae) 

(Barros et al., 2004). Part of the feeding 

strategy for this population may be to 

induce shoaling of fish by physically 

corralling them, allowing individuals to 

forage and feed successfully, even 

within murky nearshore waters (Sheehy, 

2009). In general, the prey species of the 

humpback dolphin include small fish 

which are generally not commercially 

valuable to local fisheries (Barros et al., 

2004; Sheehy, 2009). 

Little is known about the life history 

and reproduction of ETS humpback 

dolphin. In some cases, comparison of 

the ETS population with other 

populations may be appropriate, but one 

needs to be cautious about making these 

comparisons, as environmental factors 

such as food availability and habitat 

status may affect important rates of 

reproduction and generation time in 

different populations. A recent analysis 

of life history patterns for individuals in 

the Pearl River Estuary (PRE) population 

is the best proxy for the ETS population. 

Like the ETS population, the PRE 

population inhabits estuarine and 

freshwater-influenced environments in 

similar proximity to anthropogenic 

activity (Jefferson et al., 2012). 

Maximum longevity for the PRE 

population is estimated to be greater 

than 38 years (Jefferson et al., 2012). 

Evidence from multi-year photoanalysis of the ETS population 

demonstrated that adult survivorship is 

high, 0.985, suggesting that this 

population also has a relatively long 

lifespan (Wang et al., 2012). In general, 

it is inferred that the population has 

long calving intervals, between 3 and 5 

years (Jefferson et al., 2012). Gestation 

lasts 10–12 months (Jefferson et al., 

2012). Weaning may take up to 2 years, 

and strong female-calf association may 

last 3–4 years (Karczmarski et al., 1997; 

Karczmarski, 1999). Peak calving 

activity most likely occurs in the 

warmer months, but exact peak of 

calving time may vary geographically 

(Jefferson et al., 2012). Age at sexual 

maturity is late, estimated at between 12 

and 14 years (Jefferson et al., 2012). 

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DPS Analysis 

The following section provides our 

analysis, based on the best available 

science and the DPS Policy, to 

determine whether the ETS humpback 

dolphin population qualifies as a DPS of 

the taxon. 

Discreteness 

The Services’ joint DPS Policy states 

that a population segment of a 

vertebrate species may be considered 

discrete if it satisfies either one of the 

following conditions: (1) It is markedly 

separated from other populations of the 

same taxon as a consequence of 

physical, physiological, ecological, or 

behavioral factors (quantitative 

measures of genetic or morphological 

discontinuity may provide evidence of 

this separation); or (2) it is delimited by 

international governmental boundaries 

within which differences in control of 

exploitation, management of habitat, 

conservation status, or regulatory 

mechanisms exist that are significant in 

light of section 4(a)(1)(D) of the ESA (61 

FR 4722; February 7, 1996). 

Individuals from the ETS population 

exhibit pigmentation that differs 

significantly from nearby populations 

along the mainland coast of China, and 

evidence suggests that pigmentation 

varies geographically across the species’ 

range (Jefferson et al., 2001; Jefferson et 

al., 2004a; Wang et al., 2008). Across the 

species, pigmentation changes as 

individuals mature. When young, 

dolphins appear dark grey with no or 

few light-colored spots; as they age, they 

transform to mostly white (appearing 

pinkish), as dark spots decrease with 

age. In particular, the developmental 

transformation of pigment differs 

significantly between ETS and nearby 

Chinese humpback dolphin 

populations; specifically, the spotting 

intensity (density of spots) on the dorsal 

fin of the ETS population is 

significantly greater than that of four 

mainland Chinese populations, 

including the other nearby populations 

in the Pearl River Estuary and Jiulong 

River estuaries (Wang et al., 2008). 

Significantly greater spotting intensity 

on the dorsal fin of the ETS population 

is consistent, regardless of age (Wang et 

al., 2008). Further, the ETS humpback 

dolphin never loses the dark dorsal fin 

spots completely, as has been observed 

in older individuals of other humpback 

dolphin populations (Wang et al., 2008). 

In contrast, dorsal fins of Chinese 

populations are strikingly devoid of 

spots, compared to their bodies, 

throughout most of their lives, except 

when they are very young or very old 

(Wang et al., 2008). These differences in 

pigmentation can be used to reliably 

differentiate between the ETS 

humpback dolphin and nearby Chinese 

populations (Wang et al., 2008). Thus, 

we consider these significant differences 

in pigmentation of the ETS humpback 

dolphin as evidence of its discreteness. 

Several researchers have suggested 

that the ETS population of the 

humpback dolphin is physically and 

geographically isolated from other 

populations, based on the fact that 

individuals have not been observed 

crossing or to have crossed the Strait of 

Taiwan, despite repeated surveys of 

Chinese and Taiwanese populations 

using photo-identification techniques 

(Wang et al., 2004; Wang et al., 2007b; 

Chen et al., 2010; Wang et al., 2011; 

Wang et al., 2012). For instance, a 

detailed analysis of more than 450 

individually-recognizable dolphins 

catalogued for Taiwanese and Chinese 

populations revealed no matches among 

them (Wang et al., 2008). Movement of 

Sousa chinensis is thought to be limited 

to shallow water and nearshore habitat 

(Karczmarski et al., 1997; Hung et al., 

2004). Water depth and fast-moving 

currents within the Eastern Taiwan 

Strait are thought to isolate the ETS 

population from Chinese populations, 

despite their relatively close geographic 

proximity (Wang et al., 2004; Wang et 

al., 2008; Wang et al., 2011; Wee et al., 

2011; Wang et al., 2012). In fact, the ETS 

population has never been observed in 

waters greater than 30 meters depth 

(Wang et al., 2007b). Evidence suggests 

that the ETS population of the 

humpback dolphin has a narrow home 

range, and does not migrate seasonally 

or mix with Chinese populations (Wang 

et al., 2011). The population has been 

shown to inhabit the shallow, narrow 

habitat on the western coast of Taiwan 

throughout the year, and exhibits strong 

site fidelity (Wang et al., 2011). 

The evidence for geographic isolation 

is based on limited survey data 

collected since 2002, which focused 

only on nearshore waters at certain 

times of year and did not survey the 

Strait waters between mainland China 

and Western Taiwan (Wang et al., 2004; 

Wang et al., 2011; Wang et al., 2012). 

Thus, the possibility for Indo-Pacific 

humpback dolphin migration or 

emigration across the Strait cannot be 

eliminated entirely. However, the best 

available scientific information 

indicates that the species is found 

primarily in shallow nearshore habitat, 

and the ETS population has never been 

observed in waters greater than 30 

meters, and thus migration or 

emigration across the deeper Strait is 

thought to occur rarely, if ever. 

The best available data suggest that 

the ETS humpback dolphin population 

is discrete from all other populations of 

the species based on its morphological 

differences. Although limited, the best 

available data also suggest that the ETS 

humpback dolphin population is 

geographically isolated from other 

populations. The morphological 

differences and geographic isolation set 

this population apart from other 

populations of the Indo-Pacific 

humpback dolphin, and thus, we 

conclude that the ETS humpback 

dolphin population meets the 

discreteness criterion of the DPS Policy. 

Significance 

When the discreteness criterion is met 

for a potential DPS, as it is for the ETS 

humpback dolphin population, the 

second element that must be considered 

under the DPS Policy is the significance 

of the DPS to the taxon as a whole. 

Significance is evaluated in terms of the 

importance of the population segment to 

the taxon to which it belongs, in this 

case the species Sousa chinensis. Some 

of the considerations that can be used 

under the DPS Policy to determine a 

discrete population segment’s 

significance to the taxon as a whole 

include: (1) Persistence of the 

population segment in an unusual or 

unique ecological setting; (2) evidence 

that loss of the population segment 

would result in a significant gap in the 

range of the taxon; and (3) evidence that 

the population segment differs markedly 

from other populations of the species in 

its genetic characteristics. 

The ETS humpback dolphin 

population occurs in an ecological 

setting similar to populations occurring 

along the coast of mainland China, and 

many features of its habitat and ecology 

are similar to those of populations 

throughout the range of the species, as 

discussed above. Throughout its range, 

the Indo-Pacific humpback dolphin is 

consistently associated with coastal 

river output and is found in shallow 

nearshore waters (Jefferson et al., 2001). 

It displays no apparent preference for 

clear or turbid waters (Karczmarski et 

al., 2000). The habitat and ecosystem 

use of the species differ in some ways 

geographically, but evidence suggests 

that the dolphin is an opportunistic 

piscivore, and thus does not exhibit 

unique or restricted feeding ecology 

across its range (Jefferson et al., 2001). 

In Thailand, Malaysia, and Indonesia, 

the species occurs in tropical seagrass, 

coral, and mangrove lagoons not present 

in ETS humpback dolphin habitat 

(Beasley et al., 1997; Smith et al., 2003; 

Adulyanukosol et al., 2006; 

Jaroensutasinee et al., 2011; Chersukjai 

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et al., 2013). In India, the species is 

associated with nearshore habitat 

consisting of mangroves, corals, and 

tidal mudflat, heavily influenced by 

monsoons that regulate the influx of 

freshwater to the system (Sutaria et al., 

2004). The ETS humpback dolphin 

habitat is most similar to that of coastal 

Chinese populations, with more 

temperate water, soft muddy substrate, 

and consistent input from river systems. 

The ETS humpback dolphin habitat 

differs from the habitat occupied by 

mainland Chinese populations in some 

ways, with nearby rivers generally 

smaller than those in mainland China, 

and with warmer waters in the winter 

due to the influence of the Kuroshio 

Current, which periodically moves into 

the Strait of Taiwan (Chern et al., 1990; 

Jan et al., 2002; Wang et al., 2008). 

However, feeding ecology, prey 

availability, and prey preference are 

thought to be similar in mainland China 

and Taiwan (Barros et al., 2004; Wang 

et al., 2007a), so these small differences 

in habitat do not seem to have 

significant effects on the species’ 

ecology. 

The presumed habitat of the ETS 

humpback dolphin is narrower in 

offshore width than that of other studied 

populations of the taxon. For instance, 

the ETS population is thought to inhabit 

a small area of coastal shallow waters 

within 3 km from the shore (Wang et al., 

2007b). In contrast, Chinese populations 

inhabit a broader shallow area ranging 

tens of kilometers offshore, where 

dolphins can range farther from the 

coastline without moving into deeper 

water (Hung et al., 2004; Chen et al., 

2011). While the ETS population 

exhibits some behavioral differences, 

such as increased cooperative calfrearing and social connectivity, as 

compared to Chinese populations 

(Dungan et al., 2011), it is uncertain 

whether or not these differences are 

adaptive or facultative, and simply 

based on the population’s low 

abundance. Thus, insufficient evidence 

exists to suggest significant differences 

in the dolphin’s ecology or adaptation 

have derived from the differences in the 

physical parameters of its environment. 

Therefore, differences in the habitat and 

ecological setting of the ETS humpback 

dolphin do not set it apart from the rest 

of the taxon, and do not appear to relate 

to significant selection pressures 

affecting the population’s foraging, 

behavior, or ecology. 

There is no evidence to suggest that 

loss of the ETS humpback dolphin 

population would result in a significant 

gap in the range of the taxon. The ETS 

humpback dolphin population 

constitutes a small and peripheral 

portion of the entire range of the 

species, and its loss would not inhibit 

population movement or gene flow 

among other populations of the species 

(Lin et al., 2012). The ETS humpback 

dolphin is distributed throughout only 

512 square kilometers of coastal waters 

off Western Taiwan; this small range is 

not geographically significant in 

comparison to the taxon’s range 

throughout the coastal Indo-Pacific and 

Indian Oceans. 

There are no data to show that the 

genetic characteristics of the ETS 

humpback dolphin population differ 

markedly from other populations in a 

significant way. While pigmentation of 

the ETS population is significantly 

different from other populations within 

the taxon (Wang et al., 2008), whether 

the pattern is adaptive or has genetic 

underpinnings is unknown. In other 

cetacean species, differences in 

pigmentation have been hypothesized to 

relate to several adaptive responses, 

allowing individuals to hide from 

predators, communicate with 

conspecifics (promoting group 

cohesion), and disorient and corral prey 

(Caro et al., 2011). However, the 

differences in ETS humpback dolphin 

pigmentation may be a result of a 

genetic bottleneck from the small size of 

this population (less than 100 

individuals) and the possibility that it 

represents a single social and/or family 

group. Such small populations are more 

heavily influenced by genetic drift than 

large populations (Frankham, 1996). 

Insufficient data exist to determine 

whether significant differences in ETS 

humpback dolphin pigmentation relate 

to the functional divergence of the 

population, or are simply a product of 

genetic drift and a genetic bottleneck. 

The best data available thus lead us to 

conclude that loss of the ETS humpback 

dolphin population would not result in 

significant loss of overall genetic or 

ecological diversity of the taxon as a 

whole. 

DPS Conclusion and Proposed 

Determination 

According to our analysis, the ETS 

humpback dolphin population is 

considered discrete based on its unique 

pigmentation patterns, which set it apart 

morphologically from the rest of the 

taxon, and evidence for its geographic 

isolation. However, while discrete, the 

ETS humpback dolphin population does 

not meet any criteria for significance to 

the taxon as a whole. The ecological 

setting it occupies is similar to that of 

the rest of the species, loss of the 

population would not constitute a 

significant gap in the taxon’s extensive 

range, and no genetic or other data have 

demonstrated that the population makes 

a significant contribution to the 

adaptive, ecological, or genetic diversity 

of the taxon. As such, based on the best 

available data, we conclude that the ETS 

humpback dolphin population is not a 

DPS and thus does not qualify for listing 

under the ESA. This is a final action, 

and, therefore, we do not solicit 

comments on it. 

Dusky Sea Snake 

The section below presents our 

analysis of the status of the dusky sea 

snake, Aipysurus fuscus. Further details 

can be found in Manning (2014). 

Species Description 

The dusky sea snake, Aipysurus 

fuscus, is a species within the family 

Elapidae, which is a very diverse family 

of venomous snakes. The genus 

Aipysurus contains seven species, six of 

which are restricted to Australasian 

waters. The dusky sea snake is brown, 

blackish-brown, or purplish-brown with 

wide ventral scales and diamondshaped body scales that are smooth and 

imbricate (i.e., overlapping). There are 

generally 19 scale rows around the neck, 

19 around the mid-body, and 155 to 180 

ventral scales (Rasmussen, 2000). The 

dusky sea snake is completely aquatic 

and, like all sea snakes, has a paddlelike tail for swimming. Its maximum 

total length is about 90 cm (Rasmussen, 

2000). Growth rates for the dusky sea 

snake have not been documented, but 

reported growth rates for other sea 

snakes range from 0.07–1.0 mm per day 

and decline with age (Heatwole, 1997). 

The maximum lifespan for dusky sea 

snakes has been assumed to be about 10 

years, and age at first maturity has been 

assumed to be about 3–4 years 

(Lukoschek et al., 2010). Generation 

length is thought to be approximately 5 

years (Lukoschek et al., 2010). 

Despite its aquatic existence, and like 

all reptiles, the dusky sea snake lacks 

gills and must surface to breathe air. 

Dive durations vary by species, but most 

sea snakes typically stay submerged for 

about 30 minutes, and some for up to 

1.5–2.5 hours (Heatwole and Seymour, 

1975). Maximum dive depth for dusky 

sea snakes is unknown, but co-occurring 

members of this genus are considered 

‘‘shallow’’ and ‘‘intermediate’’ depth 

species that dive no deeper than 20 m 

or 50 m, respectively (Heatwole and 

Seymour, 1975). 

The dusky sea snake is viviparous, 

meaning embryos develop internally 

and young undergo live birth. Because 

this species never ventures on land, 

mating occurs at sea and young are born 

alive in the water. Within the genus 

Aipysurus, the number of young per 

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brood is small, usually less than four, 

and young are relatively large at birth 

(Cogger, 1975). Timing and seasonality 

of the dusky sea snake’s breeding cycles 

are unknown, and very little is known 

about the juvenile life stage. 

The dusky sea snake preys mainly on 

labrid (e.g., wrasses) and gobiid (e.g., 

gobies) fishes, and to a lesser extent, fish 

eggs (McCosker, 1975). Food 

competition among sympatric sea 

snakes is thought to be minimal, based 

on examinations of diet composition for 

sympatric sea snakes (McCosker, 1975; 

Voris and Voris, 1983). Feeding 

behavior of dusky sea snakes has not 

been thoroughly investigated; however, 

during surveys at Ashmore Reef, 

Australia, Guinea and Whiting (2005) 

commonly saw dusky sea snakes over 

sand bottom habitat and watched one 

snake actually force its head and about 

15 percent of its body into the sand. 

However, because it emerged without a 

prey item (Guinea and Whiting, 2005), 

it is unclear whether this was foraging 

or some other behavior. Like their 

terrestrial relatives, sea snakes swallow 

their prey whole and therefore must 

have some strategy for subduing them. 

McCosker (1975) hypothesized that the 

highly toxic venom of sea snakes is 

probably more of a feeding adaptation 

than a defensive one. 

The dusky sea snake is a benthic, 

coral reef-associated species endemic to 

several shallow emergent reefs of the 

Sahul Shelf off the coast of Western 

Australia in the Timor Sea, between 

Timor and Australia. These reefs are 

relatively isolated and lie at the edge of 

the continental shelf over several 

hundred kilometers from the mainland. 

The dusky sea snake has been reported 

to occur at Ashmore, Scott, 

Seringapatam, and Hibernia Reefs and 

Cartier Island; however, individual 

surveys have not consistently recorded 

dusky sea snakes at all of these 

locations. For example, in transect 

surveys conducted by Minton and 

Heatwole (1975) over several weeks 

during December 1972 and January 1973 

at Ashmore, Scott, and Hibernia Reefs 

and Cartier Island, dusky sea snakes 

were recorded at Scott and Ashmore 

reefs only. Extensive surveys conducted 

more recently at Ashmore Reef, where 

dusky sea snakes were once relatively 

common, have located no specimens 

(Guinea, 2013; Lukoschek et al., 2013). 

Lukoschek et al. (2010) estimated that 

the area of occurrence of dusky sea 

snakes is probably less than 500 km2. 

During their surveys, Minton and 

Heatwole (1975) observed dusky sea 

snakes in shallow water (<10 m) as well 

as in the 12 to 25 m depth-zone. They 

were observed in areas of moderate to 

heavy coral growth, but they were also 

observed to congregate in sandybottomed gullies and channels (Minton 

and Heatwole, 1975). Home-range size 

and site fidelity of individual dusky sea 

snakes has not been evaluated. 

However, a short-term (6–9 days), 

telemetry study on the sympatric olive 

sea snakes (A. laevis) and a long-term (8- 

year), mark-recapture study on the 

turtle-headed sea snake 

(Emydocephalus annulatus) suggest that 

home-ranges of sea snakes are small, 

movement of adults is very limited, and 

longer-distance dispersal may be due 

mainly to passive transport, such as by 

currents and storms (Burns and 

Heatwole, 1998; Lukoschek and Shine 

2012). While it is very plausible that 

adult A. fuscus are similar to these other 

species, research to evaluate adult and 

juvenile A. fuscus habitat use and 

movement is needed. 

Sea snakes typically have patchy 

distributions and can be found in very 

dense aggregations in certain locations 

within their ranges (Heatwole, 1997). 

This patchiness complicates efforts to 

understand habitat use patterns, as 

seemingly suitable habitat can remain 

unoccupied. On a smaller spatial scale, 

distributions of sea snake fauna on 

Australian reefs appear to be influenced 

by water depth, substrate type, and 

feeding strategies (McCosker, 1975; 

Heatwole, 1975b). Other biotic factors, 

such as limited juvenile dispersal, may 

also contribute to the observed patchy 

distributions (Lukoschek et al., 2007a). 

Overall, however, causative factors for 

observed distributions are not 

completely understood. 

Population Abundance, Distribution, 

and Structure 

There are no historical or current 

population estimates for the dusky sea 

snake. However, multiple reefs have 

been surveyed repeatedly, and although 

survey methodologies have varied, the 

data provide some indication of 

population trends for some locations. 

For Ashmore Reef in particular, the 

survey data provide a strong indication 

of severe population decline and 

possible extirpation. Older surveys 

dating from 1972 to 2002 by various 

researchers indicate that the relative 

abundance of A. fuscus was fairly 

consistent and represented about 10–23 

percent of the sea snakes observed (see 

Table 1, Manning, 2014). A footnote in 

Smith (1926) also indicates that a 

sample of 27 dusky sea snakes (out of 

an ∼100-specimen sea snake collection) 

had recently been collected for him at 

Ashmore Reef. The dusky sea snake, 

however, has not been recorded in a 

single survey conducted at Ashmore 

Reef after 2005, despite considerable 

effort (Lukoschek et al., 2013; Table 1, 

Manning, 2014). Based on reef area data 

reported in Skewes et al. (1999), 

Ashmore Reef represents about 40 

percent of the dusky sea snake’s 

historical reef habitat. Extirpation from 

this reef would represent a substantial 

change in the species’ distribution and 

abundance. 

A survey in 2005 at Hibernia Reef 

indicated a relatively low abundance of 

A. fuscus, and the most recent surveys, 

conducted in 2012 and 2013, have failed 

to detect any dusky sea snakes despite 

extensive survey effort (Guinea, 2005; 

Guinea, 2013). Dusky sea snakes were 

observed in surveys conducted at Scott 

Reef in 1972/73, 2006, 2012 and 2013; 

however, their relative abundance varies 

across the surveys, and no trends are 

detectable given the limited data (see 

Table 1, Manning, 2014). For example, 

Guinea (2012) visited Scott Reef in 

February, 2006, and reported that dusky 

sea snakes, as the third-most abundant 

species, made up 15 percent of the total 

sea snake sightings (Guinea, 2013). 

Portions of Scott Reef were surveyed 

again in 2012 and 2013, and dusky sea 

snakes made up only 3.2 percent and 

7.4 percent of the total sightings 

respectively for each year (Guinea, 

2013). At Seringapatam Reef and Cartier 

Island, A. fuscus is rare or potentially 

absent. Overall, while these limited 

abundance data are very difficult to 

interpret, they indicate that dusky sea 

snakes have not been present in high 

numbers in any recent reef surveys 

(Table 1, Manning, 2014). 

The dusky sea snake has a restricted 

range, and structure and connectivity of 

populations is uncertain. Assuming that 

A. fuscus is extirpated from Ashmore 

Reef, Sanders et al. (2014) recently 

estimated that the dusky sea snake’s 

range is now less than 262 sq km. 

Although structure and connectivity of 

reef populations of A. fuscus have not 

been studied directly, some information 

may be gleaned from several studies on 

the olive sea snake, A. laevis, a 

sympatric congener that is larger in size, 

more common, and more widely 

distributed than A. fuscus, but is very 

closely related to A. fuscus (Sanders et 

al., 2013b). As mentioned above, a 

short-term (6–9 days) tracking study on 

A. laevis suggests that adults of this 

species have small home ranges (1,500– 

1,800 sq m) and undergo limited active 

dispersal (Burns and Heatwole, 1998). 

Results of that study are somewhat 

supported by analyses by Lukoschek et 

al. (2007b) of mitochondrial DNA 

(mtDNA) from 354 olive sea snakes 

collected across its range, including 

some samples from Hibernia, Scott, and 

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Ashmore reefs and Cartier Island. Based 

on their results, Lukoschek et al. 

(2007b) concluded that gene flow among 

the reefs of the Timor Sea is low, and 

that olive sea snakes at these reefs have 

been diverging for some time. A 

subsequent analysis of microsatellite 

DNA from the same specimens indicates 

that two of the most distant Timor reef 

populations of A. laevis are significantly 

diverged (Lukoschek et al., 2008). 

However, the degrees of divergence of 

other reef populations were not 

statistically significant, and there was 

no clear isolation-by-distance 

relationship (Lukoschek et al., 2008). 

Although not conclusive, the available 

information for the olive sea snake and 

the fact that dusky sea snakes also lack 

a dispersive larval phase, suggest 

connectivity of A. fuscus may be limited 

among some reefs within the region. 

Limited inter-population exchange 

would increase the extinction risk and 

reduce the recovery potential for local 

populations that have experienced 

severe declines or have been lost. 

Summary of Factors Affecting the Dusky 

Sea Snake 

Available information regarding 

current, historical, and potential threats 

to the dusky sea snake was thoroughly 

reviewed (Manning, 2014). Although 

causes for observed declines in dusky 

sea snake have not been conclusively 

determined, we found that the species is 

being threatened by hybridization. 

Other possible threats include vessels, 

pollution, climate change, and 

inadequate regulatory mechanisms. We 

summarize information regarding each 

of these threats below according to the 

factors specified in section 4(a)(1) of the 

ESA. Available information does not 

indicate that disease, predation, or 

overutilization (including bycatch) are 

operative threats on this species; 

therefore, we do not discuss those 

further here. See Manning (2014) for 

additional discussion of all ESA Section 

4(a)(1) threat categories. 

The Present or Threatened Destruction, 

Modification, or Curtailment of Its 

Habitat or Range 

Aipysurus fuscus is dependent on 

coral reefs for prey and shelter, and loss 

of live coral is a possible mechanism 

contributing to the decline of A. fuscus 

at locations such as Ashmore Reef. Coral 

reefs of the Sahul Shelf experienced 

widespread bleaching in response to El 

Nin ̃ o events in 1998 and 2003. Ashmore 

Reef experienced bleaching in 1998 and 

again, to an apparently greater extent, in 

2003 (Lukoschek et al., 2013). However, 

because there are no estimates of coral 

coverage prior to 1998, the extent of 

coral loss following the 1998 event has 

not been quantified. Widespread 

mortality of corals was documented in 

response to the 2003 bleaching event, 

and average live coral coverage was 

reduced to 10 percent (Kospartov et al., 

2006; as cited in Lukoschek et al., 2013). 

Surveys conducted in 2005 and 2009 

indicated that recovery of corals at 

Ashmore Reef was rapid but delayed by 

about 7 years (Ceccarelli et al., 2011). 

Overall, there has been an eight-fold 

increase in hard coral coverage from 

1998 to 2009 (Hale and Butcher, 2013), 

with all of the recorded recovery 

occurring after 2005. Meanwhile, survey 

data suggest complete loss of dusky sea 

snakes at Ashmore Reef after 2005. 

Existing survey data also show sharp 

declines in total sea snake abundance 

and species diversity at Ashmore Reef 

following both the 1998 and 2003 

bleaching events (Lukoschek et al., 

2013). These patterns are consistent 

with a hypothesis that loss of live corals 

affects reef-associated sea snakes. 

The patterns at Ashmore Reef are 

contrasted, however, by those observed 

at Scott Reef. Following the 1998 

bleaching event, a greater than 80 

percent loss of hard and soft coral cover 

occurred, which translated into a 

reduction of live coral coverage to a 

total of roughly 10 percent (Smith et al., 

2008). The 1998 El Nin ̃ o event 

represents the most extreme 

temperature anomaly recorded for Scott 

Reef, and involved a rapid rise in water 

temperatures that remained above 

normal for two months (NOAA, 2013). 

Almost 6 years after the bleaching event 

(in 2004), the hard corals had partially 

recovered to 40 percent of their prebleaching cover, the soft corals showed 

no sign of recovery, and community 

composition of corals remained 

significantly altered (Smith et al., 2008). 

Within 12 years after the event (by 

2010), coral cover, recruitment, 

community composition, and generic 

diversity were similar to pre-bleaching 

years (Gilmour et al., 2010). Several 

lesser disturbances, including two 

cyclones and the 2003 El Nin ̃ o event, 

occurred during this time period and 

may have slowed the rate of recovery to 

some extent (Gilmour et al., 2013). 

Available sea snake survey data, 

spanning 1972–2013, with surveys in 

1972–73, 2006, 2012, and 2013, do not 

appear to indicate a major decline in 

abundance of dusky sea snakes at Scott 

Reef, which were relatively common 

during the surveys conducted by Guinea 

(2012) in 2006. However, the temporal 

gaps in these survey data, especially 

from 1973 to 2006, could conceal 

shorter-term patterns. 

A comprehensive understanding of 

the relationship between live coral 

cover and dusky sea snake abundance 

likely requires more detailed 

information regarding coral species 

composition, habitat complexity, and 

coral and prey fish resiliency relative to 

both the 1998 and 2003 bleaching 

events. Such an analysis might offer 

further insights into the differing 

response patterns at the two reefs, and 

an indication of whether sea snake 

abundance is driven by live coral 

coverage over timescales relevant to 

these disturbances. At this time, 

however, because a clear or consistent 

pattern does not emerge from the 

available data regarding dusky sea snake 

abundances at Ashmore and Scott reefs 

in relationship to these two bleaching 

events, we cannot conclude that loss of 

live coral is contributing to the decline 

of the dusky sea snake. 

The reefs where dusky sea snakes are 

found lie more than several hundred 

kilometers offshore and thus enjoy a 

considerable degree of protection from 

human activities and land-based sources 

of pollution. Despite this remoteness, 

the reefs may experience some 

degradation as a result of vessel traffic. 

Anchor damage, pollution from 

contaminated bilge water, and marine 

debris are among the potential issues 

identified at Ashmore Reef, which 

experiences a relatively high level of 

traffic from Indonesian fishers, yachts, 

merchant ships, and illegal entry vessels 

(Whiting, 2000; Lukoschek et al., 2013). 

The mechanisms for and extent to 

which these boat-based habitat threats 

are impacting dusky or any other sea 

snake species of the Timor Sea reefs are 

unknown. 

The extensive oil and gas industry 

activity in this region may also be a 

possible source of disturbance affecting 

dusky sea snakes and their habitat. 

Exploration and extraction activities 

within the Ashmore Platform began in 

1968 (Geoscience Australia, 2012) and 

are expected to continue for some time, 

given the significant resources within 

this region. Ashmore Reef and Cartier 

Islands lie about 50–80 km west of the 

main offshore wells in the Timor Sea, 

and the closest exploration wells are 36 

km away (Russell et al., 2004). However, 

Scott Reef lies directly above a 

significant portion of the Torosa 

Reservoir, where drilling for natural gas 

is expected to start by 2017. The 

development of the natural gas facility 

in this area will mean increased vessel 

traffic and potentially light, sound, and 

chemical pollution. The area is also 

expected to experience minor 

subsidence or compaction as the gas is 

removed (Woodside Energy LTD, 2013). 

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Whether, and the degree to which, any 

of these threats or a combination of 

these threats will impact dusky sea 

snakes is not yet known. 

Unfortunately, extremely limited 

information also exists regarding the 

toxic effects of oil exposure on sea 

snakes. Oil spills, which occur more 

frequently as a result of vessel or 

pipeline incidents rather than 

exploration and drilling activities 

(www.amsa.gov.au), have also not 

occurred very often in this region. Some 

information is available from the August 

2009 explosion of the West Atlas oil rig 

on the Montara Well, which leaked oil 

and gas uncontrollably into the Timor 

Sea for 74 days until the well was 

finally capped in November 2009. 

Considered one of the worst oil-related 

spills to have ever occurred in Australia, 

the Montara leak was analogous in 

nature to the Deepwater Horizon 

disaster of April 2010 in the Gulf of 

Mexico. In an effort to rapidly assess 

impacts to multiple taxa, Watson et al. 

(2009) conducted ship-based transect 

surveys in areas around the Atlas 

drilling platform in September 2009. 

They did not observe or identify any 

dusky sea snakes; however, they did 

observe ‘‘lethargic sea snakes lying in 

thick oil (i.e., not moving much when 

approached, unable to dive)’’ and 

collected a dead horned sea snake 

(Acalyptophis peronii) from oil-affected 

waters for further analysis (Watson et 

al., 2009). The necropsy report 

indicated that this snake was in good 

physical condition, with no visible 

external or internal pathologies, and no 

oil was detected in swab samples of the 

skin (Gagnon and Rawson, 2010). 

Chemical analysis of tissues clearly 

indicated that exposure to crude oil 

occurred through ingestion of prey and 

not through inhalation (Gagnon and 

Rawson, 2010). Acalyptophis peronii is 

considered more of a diet specialist than 

the dusky sea snake and primarily 

consumes burrowing gobies (McCosker, 

1975; Voris and Voris, 1983). Because 

they saw no physical damage to the gut 

structure and no contamination of the 

tissues, Gagnon and Rawson (2010) 

concluded it was unlikely that oil 

ingestion was the primary cause of 

death. Tests for presence of chemical 

dispersants used during the spillresponse were not conducted. 

A necropsy was also performed on a 

dead sea snake landed by a commercial 

fisherman operating in the vicinity of 

the West Atlas spill on September 14, 

2009 (Gagnon, 2009). This specimen 

was identified as Hydrophis elegans, 

which is a relatively widespread and 

abundant species that preys on eels and 

other fishes (McCosker, 1975; Voris and 

Voris, 1983). The necropsy indicated 

that the snake had fed recently and that 

the stomach contents were 

contaminated with oil (Gagnon, 2009). 

Relatively high levels of polycyclic 

aromatic hydrocarbons were also 

detected in the lungs, trachea, and 

muscle tissue (Gagnon, 2009). Neither of 

two dispersant chemicals used to treat 

the spill were detected in lung samples 

(Gagnon, 2009). The necropsy report 

concluded that the likely cause of death 

for this specimen was exposure to 

petroleum hydrocarbons (Gagnon, 

2009). 

In 2012 and 2013, Guinea (2013) 

conducted surveys to evaluate the 

potential impacts of the Montara leak on 

species of marine reptiles. Potentially 

impacted areas surveyed included 

Ashmore Reef, Cartier Island, and 

Hibernia Reef; Scott and Seringapatam 

reefs were surveyed as control reefs 

(Guinea, 2013). Ashmore Reef and 

Cartier Island are 167 km west-northwest and 108 km west from the Montara 

well, respectively. Seringapatam and 

Scott reefs are several hundred km 

south-east of the Montara well and far 

from modeled oil trajectories (Guinea, 

2013). The extensive survey efforts of 

Guinea (2013) did not indicate any 

impact of the hydrocarbon release on 

marine reptiles (sea turtles and sea 

snakes) of the potentially affected reefs. 

Of the reefs surveyed, Hibernia Reef and 

Cartier Island had the highest sea snake 

density; however, no sea snakes were 

observed at Ashmore Reef, where sea 

snake abundance and diversity had 

already declined to very low levels prior 

to the 2009 incident (Guinea, 2013). 

Overall, these data suggest that while 

there are likely to be acute impacts to 

sea snakes in response to major spills, 

it is unlikely that pollution stemming 

from oil and gas industry activities has 

contributed to the observed declines of 

the dusky sea snake. 

Overall, based on the existing 

information, we conclude that there is a 

low likelihood that these habitat-related 

threats have contributed to the observed 

decline of the dusky sea snake. At this 

time, there is insufficient information to 

indicate whether and how the dusky sea 

snake will be affected by these habitat 

issues in the future. We do expect that 

each of the various habitat-related issues 

summarized above will continue well 

into the future, and some may worsen. 

Given that El Nin ̃ o and its associated 

warming of equatorial Pacific Ocean 

waters is a natural and reoccurring 

climate phenomenon, coral bleaching in 

response to sufficiently strong El Nin ̃ o 

events will continue. Furthermore, 

because climate warming as a 

consequence of carbon dioxide 

emissions is expected to continue (IPCC, 

2013), and elevated sea surface 

temperatures are expected to rise at an 

accelerated rate (Lough et al., 2012), loss 

of corals through bleaching events is 

expected to increase. The expansion of 

Australia’s oil and gas exploration and 

extraction in the Timor Sea may also 

result in an increased risk of oil spills 

and additional habitat threats for dusky 

sea snakes. 

Inadequacy of Existing Regulatory 

Mechanisms 

The dusky sea snake and its habitat 

receive a significant degree of regulatory 

protections. The largest potential gap in 

existing regulatory mechanism may be 

for threats related to climate change. Oil 

spills, while rare and unpredictable, and 

other oil and gas industry activities may 

also pose threats to the species as a 

consequence of inadequate management 

and regulation. We summarize the 

available information regarding related 

regulatory protections below; a more indepth discussion is available in 

Manning (2014). 

Along with all of Australia’s other 

hydrophiine sea snakes, dusky sea 

snakes are listed under the 

Commonwealth Environment Protection 

and Biodiversity Conservation Act 1999 

(EPBC Act). The EPBC Act provides a 

legal framework to protect and manage 

Australia’s nationally and 

internationally important flora, fauna, 

ecological communities, and heritage 

places that are of national 

environmental significance. Under the 

EPBC Act, no one may ‘‘kill, injure, 

take, trade, keep or move a member of 

a native species’’ within any reserve 

without a permit (Commonwealth of 

Australia, 2000). The EPBC Act requires 

that surveys be conducted for listed 

marine species. The EPBC Act also 

provides that the Australian 

Government Minister for the 

Environment may make or adopt a 

recovery plan for a listed species, to set 

out the research and management 

actions needed to stop the decline of the 

species and support its recovery. There 

are no recovery plans in place for any 

sea snake species, however 

(www.environment.gov.au/topics/

biodiversity/threatened-speciesecological-communities/recovery-plans). 

Thus, while the dusky snake receives 

substantial protection under the EPBC 

Act, without a recovery plan, that 

protection may not be enough to help 

stabilize and recover the species. 

Two of the five main reefs within the 

dusky sea snake’s historical range, 

Ashmore Reef and Cartier Island, are 

protected reserves. Ashmore Reef 

National Nature Reserve was established 

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in 1983, under the National Parks and 

Wildlife Conservation Act 1975 (a 

predecessor to the EPBC Act), and later 

listed as a Ramsar Site in 2000, under 

the Ramsar Convention, which is an 

intergovernmental treaty on sustainable 

use of wetlands. In Australia, Ramsar 

Sites receive protection under the EPBC 

Act: Any action that will have or is 

likely to have a significant impact on a 

Ramsar Site requires an environmental 

assessment and approval. The EPBC Act 

also sets forth national standards for 

managing, planning, monitoring, 

involving the community in, and 

conducting environmental assessments 

of Ramsar Sites to insure consistent 

compliance with the Ramsar 

Convention. Cartier Island, a former 

British Air Force bombing range, was 

designated as a Marine Reserve in 2000. 

These two reserves cover a combined 

area of 750 km2 and are both assigned 

to IUCN category Ia—strict nature 

reserve. IUCN category Ia areas are 

protected to preserve biodiversity and 

maintain the areas for the benefit of 

scientific research. Human access to 

such areas is tightly controlled and 

limited. A small section of Ashmore 

Reserve is managed as IUCN category 

II—national park. Such areas are 

managed to protect ecosystems and 

biodiversity, and while still restricted, 

human visitation is not as limited as for 

category Ia areas. No fishing or harvest 

of any biota is allowed within the 

reserves, with the limited exception of 

finfish fishing within the category II 

area of Ashmore Reef, and then only as 

long as the fish are used for relatively 

immediate consumption. Given the lack 

of clearly identified habitat-related or 

human-disturbance-related threats to 

the dusky sea snake, there is no 

indication that these reserves and area 

protections are inadequate such that 

they have contributed to the observed 

decline of the species. 

According to the Australia 

Department of Sustainability, 

Environment, Water, Population, and 

Communities (DSEWPC) 2012 Report 

Card for marine reptiles listed under the 

EPBC Act, pollution from offshore oil 

rigs and operations is a potential 

concern for sea snakes (DSEWPC, 2012). 

This report also states that Australia has 

a strong system for regulating the oil 

and gas industry and that this system 

was strengthened further in the wake of 

the Montara oil spill. Details on how 

any particular processes or regulations 

were strengthened are not provided in 

this report and could not be found. 

Although oil spills pose a potential 

threat to the health and status of the 

dusky sea snake, oil spills are relatively 

rare, and there is insufficient 

information to indicate that the existing 

regulatory mechanisms are inadequate 

or that they have contributed to the 

decline of this species. 

Potential threats to dusky sea snakes 

stemming from anthropogenic climate 

change include elevated sea surface 

temperature, ocean acidification, and 

increased coral bleaching events (see 

below). Impacts of climate change on 

the marine environment are already 

being observed in Australia and 

elsewhere (Melillo et al., 2014; 

Poloczanska et al., 2012), and the most 

recent United Nations 

Intergovernmental Panel on Climate 

Change (IPCC) assessment provides a 

high degree of certainty that human 

sources of greenhouse gases are 

contributing to global climate change 

(IPCC, 2013). Ocean temperatures 

around Australia have increased by 0.68 

°C since 1910–1929 (Poloczanska et al., 

2012), and carbon dioxide inputs have 

lowered ocean pH by 0.1 units since 

1750 (Howard et al., 2009). Australia 

and other countries have responded to 

climate change through various 

international and national mechanisms. 

Australia signed on to the Kyoto 

Protocol in 2007 and has active 

domestic and international programs to 

lower greenhouse gas emissions 

(www.climatechange.gov.au/). However, 

in Australia, there appear to be no 

specific actions to address potential 

climate change effects on marine 

reptiles beyond monitoring (Fuentes et 

al., 2012). Because climate change 

related threats have not been clearly or 

mechanistically linked to decline of 

dusky sea snakes, the adequacy of 

existing or developing measures to 

control climate change threats is not 

possible to fully assess, nor are 

sufficient data available to determine 

what regulatory measures would be 

needed to adequately protect this 

species from climate change. While it is 

not possible to conclude that the current 

efforts have been inadequate, such that 

they have contributed to the decline of 

this species, we consider it likely that 

dusky sea snakes will be negatively 

impacted by climate change, given the 

predictions of widespread and 

potentially permanent damage to coral 

reefs in Australia (IPCC, 2013). 

Overall, we do not find there is 

substantial evidence indicating that A. 

fuscus is currently threatened by the 

lack of adequate regulatory mechanisms. 

Beyond the direct protection the species 

receives through its listing under the 

EPBC Act, the dusky sea snake receives 

additional direct and indirect protection 

within the Ashmore Reef and Cartier 

Island Marine Reserves. Given the 

predictions of worsening damage to 

coral reefs in Australia in response to 

climate change (IPCC, 2013), the largest 

potential future gap in the existing 

regulatory mechanisms appears to be 

related to climate change. 

Other Natural or Manmade Factors 

Affecting Their Continued Existence 

Elevated sea surface temperature as a 

consequence of climate change has been 

proposed as a possible threat to sea 

snakes, and we have addressed habitatrelated effects above. The IUCN Red List 

assessment for A. fuscus, suggests that 

climate-induced increases in water 

temperature may actually exceed the 

upper lethal limit for A. fuscus, and 

thereby pose a threat to the species 

(Lukoschek et al., 2010). These authors 

assumed an upper lethal limit of 36 °C, 

based on data for the pelagic sea snake, 

Pelamis platurus. Experiments to 

measure the thermal tolerances of A. 

fuscus have not been conducted. 

Sea snakes, like all reptiles, are 

ectotherms, and thus to a great extent 

are physiologically affected by 

temperature. On a large geographic 

scale, the distribution of sea snakes is 

considered to be dictated by ocean 

temperatures: Sea snakes generally do 

not occur in waters below about 18 °C 

(Davenport, 2011). Most sea snakes can 

tolerate temperatures up to a mean of 

about 39–40 °C, but tolerances may vary 

with the size of the snake and the rate 

of temperature change (Heatwole et al., 

2012). Also, although sea snakes are 

able to dive to avoid extreme 

temperatures of surface waters, they 

have limited capacity to acclimate and 

cannot thermoregulate (Heatwole et al., 

2012). 

Sea surface temperatures vary 

seasonally within the Timor Sea. The 

highest recorded oceanic water 

temperature in the Ashmore region is 31 

°C, and the highest recorded lagoon 

water temperature is 35.4 °C 

(Commonwealth of Australia, 2002). 

These temperatures are below the 

assumed upper lethal temperature limit 

for dusky sea snakes; but Australia’s 

average ocean temperatures have 

increased by over half a degree since 

1910–1929, and the rate of warming has 

accelerated since the mid-20th century 

(Poloczanska et al., 2012). Given the 

thermal tolerances of other sea snakes 

and the ocean temperatures currently 

experienced by A. fuscus at present, it 

is very unlikely that elevated ocean 

temperature has been a source of 

mortality. However, it is plausible that 

a continuation of the observed rate of 

ocean warming would, in the distant 

future, result in negative physiological 

consequences for A. fuscus. 

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Hybridization and introgression have 

recently been identified by Sanders et 

al. (2014) as a threat to the continued 

existence of A. fuscus. Hybridization, or 

the production of viable offspring 

through the crossing of genetically 

distinct taxa or groups, occurs in the 

wild for about 10 percent of animal 

species (Mallet, 2005). Hybridization 

can lead to introgression, or the 

integration of foreign genetic material 

into a genome. The conservation 

concern in this particular case is that 

reproductive barriers between the olive 

sea snake, A. laevis, and the dusky sea 

snake, A. fuscus, appear to be breaking 

down, potentially allowing A. fuscus to 

undergo reverse speciation. 

The dusky sea snake co-occurs with 

the closely-related olive sea snake 

throughout its range, and the two 

species are thought to have shared a 

common ancestor approximately 

500,000 years ago (Sanders et al., 

2013b). The olive sea snake is a 

relatively abundant and much more 

widely distributed species compared to 

the dusky sea snake. Although similar 

in appearance, the two species can be 

distinguished based on body scale rows, 

body size, and color pattern. Sanders et 

al. (2014) analyzed 11 microsatellite 

markers for A. fuscus and A. laevis 

across four reefs (Ashmore, Hibernia, 

Scott, and Seringapatam) to assess interspecific gene flow and introgression. 

Results of their genetic analyses indicate 

significant and asymmetric gene flow, 

with higher rates of introgression from 

A. laevis into the smaller A. fuscus 

population (Sanders et al., 2014). A high 

frequency of hybrids was also found at 

each of the four reefs included in the 

study area. Forty-three percent of the 

snakes sampled (n=7) at Ashmore, 55 

percent of the snakes sampled (n= 42) 

at Scott Reef, and 42 percent of the 

snakes sampled (n=12) at Seringapatam 

Reef were identified as hybrids (Sanders 

et al., 2014). At Hibernia Reef, 95 

percent of the snakes sampled (n=19) 

were hybrids (Sanders et al., 2014). 

Phenotypically, the majority of hybrids 

resembled the olive sea snake (Sanders 

et al., 2014). Whether the observed 

hybridization is a purely natural process 

or has human causes is not yet known. 

Regardless, the high rates of 

hybridization of A. fuscus with another 

species across its range may lead to the 

eventual disappearance of this 

taxonomic species and is a threat to its 

survival. 

Extinction Risk 

Although accurate and precise data 

for many demographic characteristics of 

dusky sea snakes are lacking, the best 

available data provide multiple lines of 

evidence indicating that this species 

currently faces a high risk of extinction. 

The probable extirpation of the dusky 

sea snake from Ashmore Reef, which 

constitutes about 40 percent of the 

historical reef habitat, represents a 

contraction of an already limited range 

for this species. Loss of dusky sea 

snakes from Ashmore Reef and low 

relative abundances at all other reefs, 

coupled with high rates of hybridization 

throughout the range and a presumed 

low rate of dispersal, suggest that the 

species is declining and unlikely to 

recover without intervention. The 

interaction of the threats of low and 

declining abundance, limited dispersal, 

and high rates of hybridization all 

suggest a high risk of extinction in the 

near term. 

Protective Efforts 

As mentioned previously, all of 

Australia’s hydrophiine sea snakes are 

listed and protected under the EPBC 

Act, making it illegal to kill, injure, take, 

trade, or move dusky sea snakes in 

Commonwealth waters without a permit 

(DSEWPC, 2012a). The EPBC Act also 

requires that surveys be conducted for 

listed marine species. 

Sea snakes are also identified as a 

‘‘conservation value’’ in Australia’s 

North-west Marine Bioregional Plan 

(DSEWPC, 2012b). Marine bioregional 

plans are meant to improve the way 

decisions are made under the EPBC Act, 

particularly with respect to balancing 

protection of marine biodiversity with 

the sustainable use of natural resources. 

The North-west Plan identifies activities 

that may affect sea snakes and thus 

require prior approval. National heritage 

places are also listed and protected 

under the EPBC Act. Ashmore, Scott, 

and Seringapatam reefs are all listed on 

Australia’s Commonwealth Heritage 

List, and under the EPBC Act, approval 

must be obtained before any action takes 

place that could have a significant 

impact on the national heritage values 

of these areas. 

Also mentioned previously were the 

various habitat protections currently in 

place that directly and indirectly protect 

the coral reefs within the dusky sea 

snake’s range. For example, the 

Ashmore Commonwealth Marine 

Reserve, which includes 583 km2 of 

sandy islands, coral reefs, and 

surrounding waters up to 50 m deep 

(Commonwealth of Australia, 2002), is 

almost completely closed to the general 

public. Permits may be issued to 

authorize visits for tourism or 

recreation. There are 1–2 visits per year 

by commercial tourism vessels to view 

wildlife, and about 15–20 recreational 

yachts that visit each year (Hale and 

Butcher, 2013). Indonesians have fished 

this site for centuries and subsistence 

fishing is allowed in only the IUCN 

category II portion of the reserve (Hale 

and Butcher, 2013). No commercial 

fishing is allowed in any part of the 

Reserve. The relatively pristine state of 

the site makes it attractive for the longterm monitoring and other scientific 

projects that are conducted there (Hale 

and Butcher, 2013). Starting in the late 

1980’s, Environment Australia (EA) 

contracted a private vessel and crew to 

undertake on-site management at the 

Reserve; however, as of 2000, Australian 

Customs Service took over this 

responsibility (Whiting, 2000). 

Enforcement of protections at the 

Reserve depends largely on the presence 

of Customs officials, which is not quite 

continuous (Lukoschek et al., 2013; 

Whiting, 2000). 

The Cartier Island Commonwealth 

Marine Reserve, designated in 2000 

under the EPBC Act, is completely 

closed to the public. No commercial or 

recreational fishing is allowed. General 

access and several specific activities, 

such as scientific research, photography 

and tourism, may be allowed with prior 

approval from the Director of National 

Parks issued under the EPBC Act (see 

http://www.environment.gov.au/topics/

marine/marine-reserves/north-west/

cartier-activities). 

Since the early 18th century, 

Indonesian fishers have visited and 

fished reefs within the Timor Sea, 

mainly in search of trepang, trochus, 

turtle, shark fin, and reef fishes 

(Commonwealth of Australia, 2002). In 

1974, a Memorandum of Understanding 

(MOU) was established between 

Australia and Indonesia that set out 

arrangements by which traditional 

fishers may access resources in 

Australia’s territorial sea. Because of its 

shape, the area covered by this MOU is 

often referred to as the MOU Box. The 

MOU Box, which covers an area of 

about 50,000 km2, includes the five 

main reefs where the dusky sea snake 

occurs (Skewes et al., 1999). The marine 

resources within this area are managed 

by the Australian Government, and 

traditional fishing by Indonesian fishers 

is allowed. However, as discussed 

above, certain restrictions apply within 

the Marine Reserves. Traditional 

Indonesian fishers may access parts of 

the Ashmore Reserve for shelter and 

freshwater and to visit grave sites, but, 

as mentioned previously, fishing is 

prohibited in both the Cartier Island and 

Ashmore Marine Reserves, with the 

limited exception for fishing for 

immediate consumption within the 

category II area of the Ashmore Reserve. 

There is no evidence that sea snakes 

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have been targeted by Indonesian fishers 

(Hale and Butcher, 201; Lukoschek et 

al., 2013). 

Because sea snakes are listed under 

the EPBC Act, all Australian fisheries 

are required to demonstrate that direct 

and indirect interactions with sea 

snakes are sustainable (Zhou et al., 

2012). Commercial trawls take over a 

dozen species of sea snakes (Heatwole 

1997; Wassenberg et al., 2001; Zhou et 

al., 2012), and in the absence of bycatch 

reduction devices (BRDs), an estimated 

48.5 percent of all incidentally captured 

sea snakes will die (Wassenberg et al., 

2001). BRDs are required in the prawn 

trawl fishery to minimize bycatch 

mortality and help conserve protected 

species. The only trawl fishery that 

operates within the range of the dusky 

sea snake is the North West Slope Trawl 

Fishery (NWSTF). The Australian 

Fisheries Management Authority 

(AFMA) reports that the NWSTF, which 

targets three scampi species (lobsters), is 

a low effort fishery with a very low level 

of bycatch and no documented 

interactions with threatened, 

endangered, or protected species 

(AFMA, 2012). The NWSTF is also a 

deep-water fishery, and thus unlikely to 

encounter the reef-associated dusky sea 

snake (Fry et al., 2001; Lukoschek et al., 

2007a; Lukoschek et al., 2013). As 

discussed here and in further detail in 

the status review report (Manning, 

2014), there is no indication that direct 

harvest or incidental capture poses a 

threat to the dusky sea snake. 

Sea snake products have been traded 

internationally since the 1930s (Marsh 

et al., 1994), but no sea snake species is 

currently listed under the Convention 

on International Trade in Endangered 

Species of Wild Fauna and Flora 

(CITES). Australia’s Wildlife Protection 

Act 1982 restricts the export of sea 

snake products out of Australia (Marsh 

et al., 1994). There are no data to suggest 

that the dusky sea snake is threatened 

by past, present, or future trade. 

Despite their apparent 

substantiveness, these existing and 

ongoing conservation efforts seem 

unlikely to prevent further decline of 

the dusky sea snake, because they have 

failed to prevent the decline of the 

species to date. For example, decades of 

protections at Ashmore Reef, while 

maintaining this as a relatively pristine 

reef (Hale and Butcher, 2013), have not 

prevented the severe decline and likely 

extirpation of dusky sea snakes there. 

Furthermore, the threat posed by 

hybridization is beyond the scope of 

existing protections. We are thus not 

able to conclude that the existing 

protective efforts alter the extinction 

risk for the dusky sea snake. We are not 

aware of any additional, planned or notyet-implemented conservation measures 

that would protect this species; thus, we 

did not conduct an analysis under the 

PECE. We seek additional information 

on other conservation efforts in our 

public comment process (see below). 

Proposed Determination 

Based on our consideration of the best 

available data, as summarized here and 

in Manning (2014), and protective 

efforts being made to protect the 

species, we conclude that the dusky sea 

snake, A. fuscus, is currently at high risk 

of extinction throughout its range. We 

therefore propose to list it as 

endangered under the ESA. 

Banggai Cardinalfish 

The following section describes our 

analysis of the status of the Banggai 

cardinalfish, Pterapogon kauderni. More 

details can be found in Conant (2014). 

Species Description 

The Banggai cardinalfish is a species 

within the family Apogonidae and 

genus Pterapogon. It was discovered in 

1920 by Walter Kaudern and described 

by Koumans (1933). The genus 

Pterapogon contains one other species, 

P. mirifica, from northwestern Australia 

(Allen and Donaldson, 2007). 

The Banggai cardinalfish is a 

relatively small marine fish. Adults 

generally do not exceed 55 to 57 mm 

standard length (Vagelli, 2011). The 

species is distinguished from all other 

apogonids by its tasseled first dorsal fin, 

elongated anal and second dorsal fin 

rays, and deeply forked caudal fin 

(Allen, 2000). It is brilliantly colored, 

with contrasting black and light bars 

with whitish spots over a silvery body. 

The Banggai cardinalfish has an 

exceptionally restricted natural range 

(approximately 5,500 km2) within the 

Banggai Archipelago, Indonesia. 

Populations have been introduced in 

areas of Indonesia outside of the 

Banggai Archipelago, including Luwuk 

Harbor (Bernardi and Vagelli, 2004), 

Palu Bay (Moore and Ndobe, 2007), 

Lembeh Strait (Erdmann and Vagelli, 

2001), Tumbak (Ndobe and Moore, 

2005), Kendari Bay (Moore et al., 2011), 

and north Bali (Lilley, 2008). These 

introductions are a result of discards 

from the ornamental live reef aquarium 

trade and introductions by dive-resort 

operators to support the tourist industry 

(Vagelli, 2011). The introduced 

populations are an artifact of the 

commercial ornamental live reef trade 

and are not part of any conservation 

program to benefit the native 

populations. Because we interpret the 

ESA as conserving species and the 

ecosystems upon which these species 

depend, we consider the natural range 

to be biologically and ecologically 

important to the species’ viability to 

persist in the face of threats. Distances 

between non-introduced populations 

range from less than 1 km (Vagelli, 

2011) up to 153 km (Vagelli et al., 2009). 

Distribution of populations is 

discontinuous, with deep water, strong 

currents, or coast exposed to severe 

weather serving as effective ecological 

barriers to migration (Bernardi and 

Vagelli, 2004; Ndobe et al., 2012; Ndobe 

and Moore, 2013). The Banggai 

cardinalfish exhibits the highest known 

degree of genetic structure of any 

marine fish (Bernardi and Vagelli, 2004; 

Hoffman et al., 2005; Vagelli et al., 

2009). Populations occurring on the 

same reef, separated by only a few 

kilometers, are genetically isolated from 

one another (Bernardi and Vagelli, 2004; 

Hoffman et al., 2005; Vagelli et al., 

2009). 

The Banggai cardinalfish is generally 

found in calm waters of sheltered bays 

or on the leeward side of islands (Allen 

and Donaldson, 2007). It inhabits a 

variety of shallow (from about 0.5 to 6 

m) habitats including coral reefs, 

seagrass beds, and less commonly, open 

areas of low branching coral and rubble. 

To avoid predators, it associates with 

microhabitats such as sea urchins and 

anemones (Vagelli, 2011). Banggai 

cardinalfish are found in waters ranging 

from 26–31 °C, but averaging 28 °C 

(Ndobe et al., 2013). 

The Banggai cardinalfish, like many 

apogonids, exhibits reversed sex roles, 

where males provide parental care and 

brood eggs in their mouths. It lacks a 

planktonic larval stage and extends the 

brooding of larvae for about 7 days after 

hatching, which results in the release of 

fully formed juveniles. Spawning occurs 

year round but peaks around September 

through October, which is a period of 

fewer storms in the region (Ndobe et al., 

2013). The Banggai cardinalfish has the 

lowest fecundity reported for any 

apogonid (Vagelli, 2011). Generation 

length (the age at which half of total 

reproductive output is achieved by an 

individual) is estimated to be 1.5 years 

(Vagelli, New Jersey Academy for 

Aquatic Sciences (NJAAS), personal 

communication cited in Allen and 

Donaldson (2007)) to 2 years (Ndobe et 

al., 2013). Its lifespan in the wild has 

been estimated at approximately 2.5–3 

years (Vagelli, 2011), with a maximum 

lifespan up to 3–5 years (Ndobe et al., 

2013). Based on a conservative estimate, 

a male could incubate/brood 

approximately 400 to 640 offspring over 

his lifespan (Vagelli, personal 

communication, 2014), of which less 

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than 5 percent may survive to adulthood 

(Vagelli 2007 as cited in CITES (2007)). 

High mortality occurs during the first 

days after release from the brood pouch 

due to predation, including parental and 

non-parental cannibalism (Vagelli, 

1999). 

Banggai cardinalfish form stable 

groups. Natural group size is difficult to 

know because group size decreases with 

fishing pressure, and most populations 

are not pristine. However, one bay 

(oyster pearl farm) in private ownership 

in the Banggai Islands had, until 2006, 

never been fished, and group size 

averaged about 13 fish, but varied from 

2–33 fish per group (Lunn and Moreau, 

2002). At the same site in 2004, group 

size varied from 1 to over 200 fish per 

group (Moore, unpublished data, 2014). 

Group size is typically less than 25 

individuals, although smaller groups are 

common and vary by age class and 

habitat type (Vagelli, 2011). 

The first scientific surveys of Banggai 

cardinalfish estimated population 

abundance and density between 1.7 

million, with a mean density of 0.03 

fishes per m2, based on a census at three 

sites in 2001 (Vagelli, 2002; Vagelli and 

Erdmann, 2002), and 2.4 million, with 

a mean density of 0.07 fishes per m2, 

based on an expanded census of 34 sites 

conducted in 2004 (CITES, 2007). In 

2007, population the density estimate of 

the expanded survey sites indicated a 

mean density of 0.08 fishes per m2

(Vagelli, 2008); however, overall 

population abundance was not reported 

for the 2007 survey. By 2011–2012, 

Ndobe et al. (in press) estimated the 

population abundance at 1.5–1.7 

million, with a mean observed density 

of 0.05 fishes per m2, reportedly for the 

24 of the 34 sites that were surveyed in 

2004 and 2007. The 2011–2012 

estimates does not include locations in 

Toado where the habitat was limited 

and density was very high (Ndobe et al., 

in press); thus, the population 

abundance estimate likely is biased low. 

However, 7 of the major sites first 

surveyed in 2004 have declined in 

abundance and mean density (Ndobe et 

al., in press), indicating the population 

has likely decreased from the 2.4 

million estimated in 2004. Although the 

mean observed density estimate of 0.03 

fishes per m2 found in the 2001 survey 

(Vagelli, 2002; Vagelli and Erdmann, 

2002) is less than the 2011–2012 survey, 

the 2001 survey was based on only three 

sites, while the 2011–2012 survey was 

based on 24 sites of the 34 sites. Ndobe 

(et al., in press) selected the expanded 

survey sites from 2004 and 2007 for 

their 2011–2012 survey based on the 

author’s previous work on habitat 

conditions and to better compare trends, 

over time, in density and abundance. 

Ndobe (et al., in press) stated that their 

2011–2012 estimate of 1.5–1.7 million 

represented 62–71 percent of the 

abundance estimate of 2.4 million from 

the 2004 survey. A total abundance 

estimate was not provided for the 2007 

survey, however mean observed density 

decreased approximately 38 percent 

between 2007 (0.08 fishes per m2) and 

2011–2012 (0.05 fishes per m2). 

Historical data on abundance are 

lacking, as surveys were done after 

harvest began in the early to mid-1990s. 

The private oyster pearl farm mentioned 

above is thought to represent a proxy for 

historical abundance by several 

researchers, though others disagree that 

the site is representative of historical 

abundance. The private oyster farm 

exists within a privately owned bay in 

Banggai Island, and fishing has been 

prohibited there since trade began, 

although illegal poaching in the bay was 

reported in 2006 (Talbot et al., 2013). 

The habitat in the bay may be similar to 

other sites that support the Banggai 

cardinalfish; thus, several researchers 

claim this population can be used as a 

proxy for a baseline of population 

abundance (Allen and Donaldson, 2007; 

Vagelli, 2008). In 2001, densities of fish 

in the private oyster pearl farm averaged 

0.63 ± 0.39 fishes per m2 (1 standard 

deviation, SD) (range: 0.28 to 1.22 fishes 

per m2) (Lunn and Moreau 2002) and 

0.58 fishes per m2 in 2004 (Vagelli 

2005). When these densities are 

compared to the densities found in the 

2001 and 2004 survey data discussed 

above, they indicate that the Banggai 

cardinalfish abundance has declined up 

to 90% from historical levels (Allen and 

Donaldson, 2007; Vagelli, 2008). 

However, several researchers (Moore, 

Sekolah Tinggi Perikanan dan Kelautan 

(STPL), personal communication 2014; 

Ndobe, Tadulako University, personal 

communication 2014) caution against 

the use of this bay as a baseline for 

population trends. Banggai cardinalfish 

population distribution is inherently 

patchy, and density is highly variable 

between and within sites of the Banggai 

Archipelago, including this bay (Moore, 

unpublished data, 2004). The 

researchers also question whether the 

habitat in the bay is comparable to other 

sites. The bay has been protected from 

degradation because it is privately 

owned and contains significant amounts 

of sheltered habitat and good quality 

microhabitat/habitat, with limited 

suitable habitat for predators of the 

cardinalfish, such as groupers and other 

larger reef fish. We acknowledge the 

debate regarding the use of the data 

from the private oyster farm as a 

baseline for historical abundance. 

However, even without that data, it is 

clear that population abundance 

estimates at sites throughout the 

Banggai Archipelago declined 

significantly between 2004 and 2011– 

2012. 

Declines and extirpations of local 

populations have been observed across 

years, likely due to directed harvest and, 

more recently, habitat destruction. In 

the 2001 survey, Bakakan Island had 

about 6,000 fish, but by the 2004 census, 

only 17 fish remained (Vagelli, 2008). In 

the 2007 survey, 350 individuals were 

found at Bakakan Island, but this was 

still well below the 6,000 fish found in 

the 2001 survey (Vagelli, 2008). In 2014, 

Moore (personal communication) 

reported that local fishers characterize 

the cardinalfish population on Bakakan 

Island as small and declining. Between 

the 2001 and 2004 surveys, the 

population density at Masoni Island 

doubled from 0.03 to 0.06 fish per m2

(an increase of approximately 150 fish 

in 3 years) (Vagelli, 2005). This increase 

is thought to have occurred in response 

to a collecting ban that the local people 

imposed in early 2003. However, in the 

2007 survey, the population was found 

to have declined to 0.008 fish per m2, 

with 38 fish recorded over the entire 

census site (the largest group consisted 

of 2 individuals). An extensive search 

around the entire island identified only 

150 fish (Vagelli, 2008). A population in 

southeast Peleng Island had 159 and 207 

fish in 2002 and 2004, respectively 

(Vagelli, 2005). However, by 2007, it 

had been practically extirpated, with 

only 27 fish found (Vagelli, 2008). 

Overharvest of microhabitat, such as 

Diadema sea urchins and sea anemones, 

and coral mining have resulted in local 

population depletions on an island off 

Liang, which was surveyed in 2004, and 

was extirpated by 2012 (Ndobe et al., 

2013). Extirpation of local populations 

has been documented in areas with 

increased harvest of microhabitat, 

combined with fishing pressure on 

Banggai cardinalfish. Interviews with 

locals and visits to several sites in 2011 

and 2012 indicate populations are 

declining in the Banggai Archipelago 

(Ndobe et al., 2013). 

Summary of Factors Affecting the 

Banggai Cardinalfish 

Next we consider whether any one or 

a combination of the five threat factors 

specified in section 4(a)(1) of the ESA 

are contributing to the extinction risk of 

the Banggai cardinalfish. We discuss 

each of the five factors below, as all 

factors pose some degree of extinction 

risk. More details are available in 

Conant (2014). 

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Present or Threatened Destruction, 

Modification, or Curtailment of Habitat 

or Range 

The illegal use of fish bombs 

(typically made with fertilizer and 

phosphorus) and cyanide to catch fish 

has resulted in significant loss of coral 

reef habitat within the Banggai 

cardinalfish range (Allen and Werner, 

2002). Damage to coral reefs due to fish 

bombs is prevalent, even in protected 

areas (Talbot et al., 2013). Cyanide is 

used to catch fish for the live reef fish 

trade, and the practice kills corals (e.g., 

see Jones and Steven, 1997; Mous et al., 

2000). Boats have degraded the coral 

reefs in the area, and clear-cutting of 

wooded slopes and mangroves has 

occurred, increasing sedimentation, 

which degrades coral reef habitat 

(Lilley, 2008). Other upland activities, 

such as agriculture and human 

population growth, have increased the 

amount of waste and nitrates in the 

marine environment, promoting algal 

blooms (Lilley, 2008), which may 

destroy coral reefs by outcompeting 

them for vital resources such as light 

and oxygen (reviewed by Fabricius, 

2005). Significant plastic, styrofoam, 

and other human-made debris occurs in 

the area (Lilley, 2008). This information 

indicates destruction of habitat is 

occurring within the Banggai 

cardinalfish’s range. Although 

quantitative data on impacts to 

cardinalfish populations are lacking, 

considerable qualitative information 

exists indicating that where habitat has 

been degraded (e.g., Tanjung Nggasuang 

and Toropot surveyed in 2004 and 2012, 

and Mbuang-Mbuang, on Bokan Island, 

surveyed in 2012), large and thriving 

Banggai cardinalfish populations spread 

over large areas can be reduced to 

isolated remnants crowded into small 

remaining patches of habitat with some 

protective microhabitat (Ndobe, 

personal communication, 2014). 

Coral reef conditions in the Central 

Sulawesi Province, including the 

Banggai Archipelago, were examined 

from 2001 through 2007 in seven 

Districts in the region (Moore and 

Ndobe, 2008). Average condition of the 

reefs was poor, and major impacts 

included coral mining, sedimentation, 

fishing, and predation (Moore and 

Ndobe, 2008). Population explosions of 

the crown-of-thorns starfish 

(Acanthaster planci), a coral predator, 

have been observed in the area, 

indicating an ecological imbalance, 

likely due to overharvest of natural 

predators and changes in hydrology and 

water quality (Moore et al., 2012). 

Surveys conducted at five sites around 

Banggai Island from 2004 through 2011 

showed coral reef cover declined by 

more than half, from 25 percent to 11 

percent (Moore et al., 2011; 2012). Major 

causes of the coral reef decline around 

Banggai Island were attributed to 

destructive fishing methods and general 

fishing pressure, coastal development, 

and the replacement of traditional 

homes with concrete and breeze-block 

dwellings, which increases the demand 

for mined coral and sand. Loss of coral 

reef cover may increase mortality of 

Banggai cardinalfish recruits due to 

cannibalism (Moore, personal 

communication, 2014; Ndobe et al., in 

press). 

Climate change may also impact 

Banggai cardinalfish habitat as a result 

of coral bleaching. Coral bleaching 

events due to warming temperatures are 

anticipated to increase by 2040 in areas 

of the Indian Ocean, including waters of 

Indonesia (van Hooidonk et al., 2013). 

Coral bleaching due to elevated water 

temperatures has not been observed 

around Banggai Island up through 

December 2011; however, extensive 

bleaching was observed in nearby 

Tomini Bay in 2010 (Moore et al., 2011; 

2012). The Banggai cardinalfish is 

restricted to shallow waters with 

ambient temperatures ranging from 28 

to 31 °C. Thus, warming temperatures 

may render habitat unsuitable, but 

specific data on impacts to the Banggai 

cardinalfish are lacking. 

Sea urchins and anemones are 

experiencing intensive and increasing 

harvest pressure, which negatively 

impacts the Banggai cardinalfish (Moore 

et al., 2012; Ndobe et al., 2012). Sea 

anemones were once abundant but were 

drastically reduced from Tinakin Laut, 

Banggai Island, which resulted in a 

collapse of the Banggai cardinalfish 

population in the area (Moore et al., 

2012). Heavy harvest of sea anemones at 

Mamboro, Palu Bay, resulted in a drastic 

reduction of new recruits and juvenile 

Banggai cardinalfish (observed since 

2006) in 2008 (Moore et al., 2011). 

Moore et al. (2011; 2012) report that 

intensive harvesting of shallow water 

invertebrates, including sea anemones 

and sea urchins, is increasing and is 

linked to socio-economic trends 

associated with consumption by local 

seaweed farmers and use as feed for 

carnivorous fish destined for the 

ornamental live reef trade. 

In addition, a disease of unknown 

origin may be damaging hard corals in 

habitat occupied by the Banggai 

cardinalfish. The disease affects the top 

sections of long-branched Acropora 

species as well as species of Porites, 

both of which are important 

microhabitat for the Banggai 

cardinalfish (Vagelli, 2011). Data are 

lacking on the extent of impact the 

disease poses to Banggai cardinalfish 

habitat. 

Overutilization for Commercial, 

Recreational, Scientific, or Educational 

Purposes 

The Banggai cardinalfish is traded 

internationally as a live marine 

ornamental reef fish. It has been 

collected in the Banggai Islands, 

Indonesia, since 1995 (Marini and 

Vagelli, 2007). The United States, 

Europe, and Asia are the major 

importers of the Banggai cardinalfish for 

the aquarium trade (CITES, 2007). The 

Banggai cardinalfish is the tenth most 

common ornamental fish imported into 

the United States (Rhyne et al., 2012). 

Banggai cardinalfish exports for the 

ornamental live reef fish trade may be 

decreasing, although systematic data are 

lacking. In 2001, up to 118,000 Banggai 

cardinalfish were sold to trade centers 

each month, with a total estimate of 

700,000–1.4 million fish traded (Lunn 

and Moreau, 2002, 2004). From 2004 

through 2006, around 600,000–700,000 

fish were traded yearly (Moore et al., 

2011). In 2008 and 2009, 236,373 and 

330,416 fish, respectively, were traded 

at Bone Bone, Toropot, and Bone Baru 

trade centers (Moore et al., 2011, 2012). 

However, these numbers do not include 

trading data from Bone Bone in 2008 

and other active centers (e.g., Panapat 

for 2008 and 2009). These collections 

centers each reported about 15,000 fish 

per month in 2007 (Vagelli, 2008; 2011). 

Vagelli (personal communication, 2014) 

estimates that 1,000,000 Banggai 

cardinalfish are currently captured each 

year for the ornamental live reef trade. 

The ornamental live reef fish trade 

has resulted in decreases in cardinalfish 

population density and extirpation of 

local populations. By 2000 (after less 

than a decade of trade), negative 

impacts on the Banggai cardinalfish 

from the trade were observed. The trade 

results in high mortality of cardinalfish 

collected. Based on interviews with 

collectors, Lilley (2008) estimated that 

only one out of every four to five fish 

collected makes it to the buyer for 

export due to high mortality and discard 

practices. Density and group size of 

cardinalfish and sea urchins are 

negatively impacted by the trade (Kolm 

and Berglund, 2003). Ndobe and Moore 

(2009) also found that populations were 

exploited, but observed high population 

density in areas where collection had 

been ongoing for some years with 

rotation between sites, indicating some 

harvest sustainability. Unfortunately, 

habitat destruction and collection and 

destruction of microhabitat (unrelated to 

the Banggai cardinalfish fishery) have 

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now greatly reduced cardinalfish 

populations at sites which had 

previously sustained periodic collection 

for more than a decade (Moore, personal 

communication, 2014). Decreases in 

population density are also evidenced 

by significant declines in the catch per 

unit effort (Vagelli, personal 

communication, 2014). In Bone Baru, 

from 1993–2000, fishers were catching 

an average of 1,000–10,000 fish per day, 

but by 2003 they only averaged 100– 

1,000 per day, with most catching 

between 200–300 fish (EC-Prep Project, 

2005). Prior to 2003, collectors from 

Bone Baru typically required one day to 

capture approximately 2,000 specimens. 

In 2007, they reported requiring one 

week to capture the same number 

(Vagelli, 2011). Vagelli (2011) reports 

similar declines for Banggai Island, 

where between 2000 and 2004, the 

reported mean catch declined from 

about 1,000 fish/hour to 25–330 fish/

hour. 

Information suggests the number of 

active participants in the trade may 

have dropped. In 2001, there were 12 

villages that collected the Banggai 

cardinalfish, but only 3 were active in 

2011 (Moore et al., 2011, 2012), and at 

least 5 villages were active in 2014 

(Moore, personal communication, 2014). 

Reported as number of collectors, the 

data indicate a decline in participation 

as well, from about 130 in 2001 (Lunn 

and Moreau, 2004) to about 80 in 2007 

(Vagelli, 2011) and 2012 (Vagelli, 

personal communication, 2014). 

In 2012, a large-scale aquaculture 

facility based in Thailand began to 

breed Banggai cardinalfish in captivity 

for export, which may alleviate some of 

the pressure to collect fish from wild 

populations (Talbot et al., 2013; Rhyne, 

Roger Williams University, unpublished 

data 2014). In 2013, approximately 

120,000 Banggai cardinalfish were 

imported into the United States from the 

Thailand facility. The volume 

represents a significant portion of 

overall United States imports of the 

cardinalfish and may even exceed the 

number of wild fish currently imported 

(Rhyne, unpublished data, 2014). Efforts 

to captive-breed the species in the 

United States are also ongoing, which 

may alleviate dependence on wildcaught cardinalfish. In the United 

States, the Florida Department of 

Agriculture and Consumer Services has 

certified eight aquaculture facilities that 

are beginning to culture and market 

farm-raised Banggai cardinalfish 

(Knickerbocker, Florida Department of 

Agriculture and Consumer Services, 

personal communication 2014). In-situ 

breeding by the fishing communities in 

the endemic area may also alleviate 

pressure on the natural population, but 

the concept requires further research 

before it can be implemented at a local 

community level (Ndobe, personal 

communication, 2014). 

Disease or Predation 

Predation and cannibalism are high 

among new recruits (Moore et al., 2012). 

However, specific data are lacking on 

whether predation pressure is 

increasing or impacting the Banggai 

cardinalfish population growth beyond 

natural levels. 

A virus known as the Banggai 

cardinalfish iridovirus (genus 

Megalocytivirus) is linked to high 

mortality of wild-caught fish imported 

for the ornamental live reef fish trade 

(Vagelli, 2008; Weber et al., 2009). The 

virus causes necrosis of spleen and 

renal tissue, which appears as darkened 

tissue. Other symptoms are lethargy and 

lack of appetite. Surveys of wild 

populations have not reported 

symptoms of the disease. Necropsies of 

over 100 fish collected in the wild and 

at holding facilities showed no 

indication of the virus (Talbot et al., 

2013). Thus, the virus is likely 

transmitted from other specimens at 

containment centers, or is carried by the 

Banggai cardinalfish and is only 

expressed as a result of stress incurred 

during the long transport process 

(Weber et al., 2009; Talbot et al., 2013) 

and may not be a concern for wild fish. 

Inadequacy of Existing Regulatory 

Mechanisms 

Current Indonesian legislation 

requires that all trade in Banggai 

cardinalfish go through quarantine 

procedures before crossing internal 

administrative borders or prior to export 

(Moore et al., 2011). Compliance 

historically has been low, but is 

improving (Moore, personal 

communication, 2014; Moore et al., 

2011). However, reported collection 

through the Fish Quarantine Data 

system, which records fish that go 

through quarantine procedures, was 

well below the total reported collection 

from Bone Baru, Toropot, and Bone 

Bone for 2008 and 2009. Bone Baru, 

Toropot, and Bone Bone reported 

collection of 236,373 fish in 2008 and 

330,416 fish in 2009. Whereas in 2008 

and 2009, the Fish Quarantine Data 

reported collection of 83,200 and 

215,950 fish, respectively (Moore et al., 

2011). Enforcement of the Fish 

Quarantine procedures is weak, and 

illegal, unregulated, and unreported 

capture and trade are still a major 

problem, especially in remote areas 

(Ndobe, personal communication, 

2014). 

Legislation is needed to establish 

fishing quotas and size limits; however, 

no legally binding regulations have been 

passed or implemented (Moore et al., 

2011). Indonesia prohibits the use of 

chemicals or explosives to catch fish 

(Fisheries Law No. 31/2004, Article 

8(1)). However, the practice continues 

(Vagelli, 2011), and damage to coral 

reefs due to fish bombs is prevalent, 

even in protected areas (Talbot et al., 

2013). 

In 2011, Indonesia had proposed to 

list the Banggai cardinalfish for 

restricted protected status under 

domestic law. But the proposal stalled 

when the Indonesian Institute for 

Science argued that the introduced 

populations meant the species was no 

longer endemic, and thus did not meet 

the criteria for protected status (Moore, 

personal communication, 2014; Ndobe, 

personal communication, 2014). In 

2007, the Banggai cardinalfish was 

proposed for listing under CITES 

Appendix II. However, the proposal 

failed. The species is listed in Annex D 

of the European Wildlife Trade 

Regulations, which only requires 

monitoring of European Union import 

levels through import notifications. 

Based on the weaknesses discussed 

above, regulatory mechanisms on the 

commercial harvest industry do not 

appear adequate to ensure the 

population will be sustainable. 

Other Natural or Manmade Factors 

Affecting Continued Existence 

Global averaged combined land and 

ocean surface temperatures show a 

warming of 0.85 °C over the period 1880 

to 2012 (IPCC, 2013). As discussed 

earlier (see Present or Threatened 

Destruction, Modification, or 

Curtailment of Habitat or Range), 

warming temperatures may destroy or 

modify habitat, but data are lacking on 

specific direct impacts to the Banggai 

cardinalfish. 

The Banggai Archipelago sits at the 

junction of three tectonic plates 

(Eurasian, Indian-Australian, and 

Pacific-Philippine Sea) and is 

vulnerable to earthquakes. An 

earthquake measuring 7.6 on the Richter 

scale occurred in 2000 and destroyed 

coral reefs in the region (Vagelli, 2011). 

Frequent earthquakes within the 

Banggai Archipelago may have 

impacted localized Banggai cardinalfish 

populations (CITES, 2007), but specific 

data are lacking. 

Extinction Risk 

The life history characteristics (i.e., 

low fecundity, high degree of parental 

care and energetic investment in 

offspring, high new recruit mortality, no 

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planktonic dispersal, high site fidelity) 

of the Banggai cardinalfish render it less 

resilient and more vulnerable to 

stochastic events than marine species 

that are able to disperse over large areas 

and recolonize sites that have been lost 

due to these events. Because the Banggai 

cardinalfish also has an exceptionally 

restricted natural range (approximately 

5,500 km2), these demographic traits 

become more important in terms of the 

extent to which the threats appreciably 

reduce the fitness of the species. The 

Banggai cardinalfish lacks dispersal 

ability and exhibits high site fidelity, 

and new recruits stay within parental 

habitat. Thus, recolonization is unlikely 

once a local population is extirpated. 

Local populations off Liang and Peleng 

Island are reported extirpated, and 

interviews with local fishermen indicate 

extirpation of small local populations 

throughout the Banggai Archipelago. 

The Banggai cardinalfish also exhibits 

high genetic population substructuring; 

thus, extirpation of local populations 

from overharvest and/or loss of habitat 

can result in loss of genetic diversity 

and further fragmentation of spatial 

distribution. In considering the 

demographic risks to the species, its 

growth rate/productivity, spatial 

structure/connectivity, and diversity are 

assigned to the high risk of extinction 

category. However, the overall 

population abundance (estimated at 1.5 

to 1.7 million) is assigned to the 

moderate risk of extinction category, 

because the abundance may allow some 

resilience against stochastic events. 

In considering the threats, we rely on 

the best available data to assess how the 

threats are currently impacting or likely 

to impact the species in the foreseeable 

future. The best available data indicate 

that several threats to the Banggai 

cardinalfish will continue and increase, 

with the species responding negatively, 

but other threats will decrease, with the 

species responding favorably. Habitat 

degradation has occurred and is 

anticipated to continue and increase in 

the foreseeable future. Although 

Indonesia prohibits the use of chemicals 

or explosives to catch fish, historically, 

compliance has been low, and data 

indicate compliance is not improving. 

Data also indicate that by 2007, harvest 

of microhabitat (sea urchins and sea 

anemones) had negatively impacted 

cardinalfish populations, and the 

harvest had increased by 2011. Moore et 

al. (2011, 2012) concluded that it would 

be difficult to establish and enforce 

local regulations for controlling the 

overharvest of microhabitat. Thus, it is 

reasonable to expect that microhabitat 

harvest will continue and increase in 

the foreseeable future, which negatively 

impacts the Banggai cardinalfish and its 

ability to avoid predators. 

Overutilization from direct harvest for 

the ornamental live reef fish trade has 

significantly impacted the Banggai 

cardinalfish and remains a concern. 

Trade continues resulting in high 

mortality, and in areas of heavy 

overexploitation, populations have been 

extirpated. However, an increase in 

compliance with the Fish Quarantine 

regulations and improved trade 

practices have occurred in recent years, 

and we anticipate compliance and trade 

practices will likely continue to 

improve in the future, which may 

mitigate impacts through sustainable 

trade. Participation in collection of 

Banggai cardinalfish for the live 

ornamental reef trade has dropped in 

recent years. Captive-bred facilities have 

recently started in the United States and 

Thailand and are anticipated to decrease 

the threat of directed harvest of the wild 

populations in the future. Predation of 

new recruits is high. Mortality from 

disease in wild-caught fish imported for 

the ornamental live reef fish trade and 

disease affecting the Banggai 

cardinalfish habitat are both plausible 

threats. However, data are lacking on 

how these threats impact the population 

and what, if any, impacts will occur and 

at what rate in the future. Climate 

change within the Banggai cardinalfish 

range will continue to affect coral reefs 

in the future, and it is reasonable to 

expect that future earthquakes that may 

destroy or modify habitat within the 

species’ range will occur at the current 

rate. 

The Banggai cardinalfish is exposed, 

and negatively responds to some degree, 

to the five threat factors discussed 

above. Although quantitative analyses 

are lacking, it is reasonable to expect 

that when these exposures are 

combined, synergistic effects may occur. 

For example, the ornamental live reef 

fish trade likely causes the expression of 

the iridovirus in the Banggai 

cardinalfish, which results in increased 

mortality. The indiscriminate harvest of 

sea anemones and sea urchins and 

destruction of coral reefs eliminates 

important cardinalfish shelter and 

substrate and increases the likelihood of 

predation. Interactions among these 

threats may lead to a higher extinction 

risk than predicted based on any 

individual threat. 

In sum, based on the life history 

characteristics of the Banggai 

cardinalfish, which indicate high 

vulnerability to demographic risks (due 

to trends in population growth/

productivity, spatial structure and 

connectivity, and diversity), coupled 

with ongoing and projected threats to 

habitat and microhabitat, commercial 

use, inadequate regulatory mechanisms, 

disease and predation, and additional 

natural or manmade factors, we 

conclude that demographic risks and 

the combination of threats to the species 

may contribute to the overall 

vulnerability and resiliency of the 

Banggai cardinalfish. The Banggai 

cardinalfish has experienced a decline 

in abundance as evidenced by the 

decrease in mean density at survey sites 

between 2004 and 2012. Moreover, at 

least some researchers believe that the 

population may have experienced a 

dramatic decline from historical 

abundance due to overharvest based on 

comparisons between populations in a 

private bay and other populations. Most 

of the species’ demographic 

characteristics put it at a high risk of 

extinction. However, the threat of 

overharvest has been and will likely 

continue to be reduced in the future. 

Further, the overall population 

abundance (1.5 to 1.7 million) may 

allow some resilience against stochastic 

events; thus, placing the Banggai 

cardinalfish at an overall moderate risk 

of extinction. 

Protective Efforts 

The Banggai cardinalfish is listed as 

‘endangered’ by the World Conservation 

Union (IUCN; Allen and Donaldson, 

2007). Although listing under the IUCN 

provides no direct conservation benefit, 

it raises awareness of the species. In 

addition, the Banggai cardinalfish was 

one of the first entrants into the Frozen 

Ark Project, which is a program to save 

the genetic material of imperiled species 

(Williams, 2004; Clarke, 2009). 

In 2007, Indonesia developed a 

national multi-stakeholder Banggai 

cardinalfish action plan (BCF–AP), 

which focused on conservation, trade, 

and management issues (Ndobe and 

Moore, 2009). As part of the BCF–AP, 

annual stakeholder meetings are held to 

share data, review progress, and set 

goals (Moore et al., 2011). The BCF–AP 

called for biophysical and socioeconomic monitoring of trade, 

population status, and habitat, and 

several organizations have begun to 

report on these activities. However, 

there is no integrated or comprehensive 

monitoring system, and long-term data 

sets are lacking (Moore et al., 2011). 

Several aspects of the BCF–AP appear to 

have improved the sustainability of the 

Banggai cardinalfish trade. Fishermen 

groups have gained legal status 

(allowing them access to various 

benefits such as funding or loan 

support), which has led to socialization 

of sustainable harvest in Bone Baru. The 

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legally-established fishermen’s group 

Kelompok BCFLestari, in Bone Baru, 

implemented collection practices 

designed to prevent capture of brooding 

males (Moore et al., 2011). Workshops 

have been held on improving capture 

methods and post-harvest care, and 

several community members have 

become active in conservation efforts. 

However, the BCF–AP officially ended 

in 2012 and so did the funding. Some 

of the stakeholders are still active and 

are likely to continue to be so, despite 

lack of government support (Moore, 

personal communication, 2014). 

As discussed earlier, compliance with 

the Fish Quarantine regulations has 

increased, which is largely due to the 

development and implementation of the 

BCF–AP (Moore et al., 2011). In 2004, 

one Banggai cardinalfish trader followed 

Fish Quarantine procedures. By 2008, 

there was a marked increase in legal 

trade, but unreported fishing still occurs 

(Moore et al., 2011). With the lapse of 

the BCF–AP, legislation is needed to 

support and restart the goals described 

in the BCF–AP, and although efforts 

have been ongoing to establish fishing 

quotas and size limits, no legally 

binding regulations have been passed or 

implemented (Moore et al., 2011). 

In 2007, the Banggai Cardinal Fish 

Centre (BCFC) was established in the 

Banggai Laut District to serve as a 

central point for sharing information 

and managing the species over a wider 

community area (Lilley, 2008; Moore et 

al., 2011). As of 2011, the BCFC had no 

electricity, no operational budget, and 

was operated on a voluntary basis 

(Moore et al., 2011). Further inhibiting 

the continued operation of the BCFC is 

that in 2013, the region was split into 

two Districts by constitutional law (UU 

No. 5/2013). The BCFC will need to be 

officially approved under the new 

District to maintain its legal status 

(Ndobe, personal communication, 

2014). 

A marine protected area (MPA) 

consisting of 10 islands was declared by 

Indonesia in 2007, with conservation of 

the Banggai cardinalfish as the primary 

goal of the Banggai and Togong Lantang 

Islands (Ndobe et al., 2012). However, 

Banggai cardinalfish populations are not 

found at Togong Lantang Island, while 

for three other islands within the 

proposed MPA with known 

populations, Banggai cardinalfish 

conservation is not included as a 

conservation goal in the designation 

(Ndobe et al., 2012). In addition, based 

on genetic analysis, only 2 of 17 known 

populations occur within the MPA, 

which led Ndobe et al. (2012) to 

conclude the MPA design was ill-suited 

for conserving the Banggai cardinalfish. 

It is uncertain whether the MPA will be 

changed in the foreseeable future to 

better suit the species. 

Although no longer active, the Marine 

Aquarium Council (MAC), an 

international non-governmental 

organization, developed a certification 

system to improve the management of 

the marine aquarium trade. MAC 

developed best practices for collectors 

and exporters, including those in 

Indonesia. Best practices include 

improvement of product quality, 

reduction in mortality rates, safer 

practices for collectors, and fairer prices 

paid to collectors. By applying the MAC 

standards, traders could be certified as 

meeting these international standards 

(Lilley, 2008). Building on the MAC 

efforts, the Yayasan Alam Indonesia 

Lestari (LINI) has worked in the Banggai 

Islands to promote a sustainable fishery 

for the Banggai cardinalfish and to 

protect habitat (Talbot et al., 2013). LINI 

focuses on surveys, capacity building, 

and training of local suppliers and reef 

restoration (Lilley, 2008). LINI’s training 

and education efforts may raise 

awareness of needed conservation 

efforts to benefit the Banggai 

cardinalfish. For example, more benign 

collection methods have been 

implemented at Bone Baru, the species 

has been adopted as a mascot, and local 

citizens craft and market items related 

to the fish. LINI is also trying to set up 

a mechanism for hobbyists to buy only 

from distributors who use best practices 

and are sustainable (Talbot et al., 2013). 

However, continued funding for the 

program is a concern (Moore, personal 

communication, 2014). 

In addition to the protective efforts 

described above, Indonesia has 

committed to develop a comprehensive 

management plan for the Banggai 

cardinalfish under the auspices of 

Indonesia’s national plan of action 

under the Coral Triangle Initiative on 

Coral Reefs, Fisheries, and Food 

Security (CTI–CFF). The CTI–CFF 

specifies a goal to use an ecosystemsbased approach to managing fisheries 

(EAFM), including a more sustainable 

trade in live reef fishes. In 2013, World 

Wide Fund for Nature (WWF), in 

partnership with STPL, implemented a 

pilot project in Central Sulawesi 

Province under the ecosystems-based 

approach and chose the Banggai 

cardinalfish as one of five fisheries case 

studies in Banggai Laut District. The 

goal is to draft local regulations for an 

EAFM for two Districts—Banggai Laut 

District (which encompasses the 

majority of the endemic Banggai 

cardinalfish populations) and Banggai 

Kepulauan District (which includes the 

Peleng Island Banggai cardinalfish 

populations). The STPL EAFM Learning 

Centre team will be implementing this 

component through January 2015. These 

efforts are likely to introduce local 

measures to sustain the Banggai 

cardinalfish trade (Moore, personal 

communication, 2014; Ndobe, personal 

communication, 2014). 

Under the PECE, conservation efforts 

not yet implemented or not yet shown 

to be effective must have certainty of 

implementation and effectiveness before 

being considered as factors decreasing 

extinction risk. The effort described 

above does not satisfy the PECE criteria 

of having a certainty of implementation 

and effectiveness. Although a pilot 

project in Central Sulawesi Province 

under the ecosystems-based approach is 

underway with the Banggai cardinalfish 

as one of five fisheries case studies, we 

lack information on how this effort will 

yield measures that will be funded, 

regulated, or regularly practiced to 

sustain the Banggai cardinalfish trade in 

the future; thus, this effort cannot be 

considered to alter the risk of extinction 

of the Banggai cardinalfish. We seek 

additional information on other 

conservation efforts in our public 

comment process (see below). 

Proposed Determination 

Based on the best available scientific 

and commercial information discussed 

above, we find that the Banggai 

cardinalfish is at a moderate risk of 

extinction, but the nature of the threats 

and demographic risks identified do not 

suggest the species is presently in 

danger of extinction, and therefore, it 

does not meet the definition of an 

endangered species. We do find, 

however, that both the species’ risk of 

extinction and the best available 

information on the extent of and trends 

in the major threats affecting this 

species (habitat destruction and 

overutilization) make it likely this 

species will become an endangered 

species within the foreseeable future 

throughout its range. We therefore 

propose to list it as threatened under the 

ESA. 

Harrisson’s Dogfish 

The following section describes our 

analysis of the status of the gulper 

shark, Harrisson’s dogfish 

(Centrophorus harrissoni). More details 

can be found in Miller (2014). 

Species Description 

Centrophorus harrissoni, or 

Harrisson’s dogfish, is a shark belonging 

to the family Centrophoridae (order 

Squaliformes). The Centrophoridae 

contain two genera: Deania (longsnouted or bird-beak dogfishes) and 

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Centrophorus, usually referred to as 

gulper sharks. ‘‘Gulper shark’’ is also the 

common name for the largest species, C. 

granulosus (White et al., 2013). 

Harrisson’s dogfish is endemic to 

subtropical and temperate waters off 

eastern Australia and neighboring 

seamounts. Specimens identified as C. 

harrissoni have also been collected 

along the Three Kings, Kermadec, and 

Norfolk Ridges north of New Zealand, 

and it has also possibly been identified 

off New Caledonia (Duffy, 2007). It is a 

demersal species, primarily found along 

the upper- to mid-continental and 

insular slopes off eastern Australia, from 

north of Evans Head in northern New 

South Wales (NSW) to Cape Hauy on 

the island of Tasmania, and on the 

Tasmantid Seamount Chain off NSW 

and southern Queensland (hereafter 

referred to as its ‘‘core range’’). It occurs 

in depths of 180 to 1000 m, with a 

principal depth range of 200 to 900 m 

(White et al., 2008; Last and Stevens, 

2009; Williams et al., 2013a). However, 

specimens have been collected in 

deeper waters from the seamounts and 

ridges north of New Zealand and off 

southeastern Australia and in shallower 

depths off eastern Bass Strait (Daley et 

al., 2002; Graham and Daley, 2011; 

Williams et al., 2013a). Gulper sharks, 

including Harrisson’s dogfish, are 

thought to conduct diel vertical feeding 

migrations, whereby the sharks ascend 

the continental slope near dusk to 

around 200 m depths to feed and then 

descend before dawn (Williams et al., 

2013a), which helps to explain the large 

depth distribution for the species. Small 

bathypelagic bony fishes (particularly 

myctophids, lantern fishes), 

cephalopods, and crustaceans have been 

found in the stomachs of C. harrissoni 

(Daley et al., 2002). 

Research studies indicate that C. 

harrissoni may also exhibit spatial 

sexual segregation (Graham and Daley, 

2011), based on the evidence that males 

tend to dominate the sex ratios on 

survey grounds and assumption that 

females must be more abundant 

elsewhere to compensate for the uneven 

sex ratios. Specifically, sex ratios varied 

from 1.5:1 to 4.9:1 along the east coast 

of Australia, illustrating the 

predominance of males (Graham and 

Daley, 2011). Two notable sites, 

however, did show females 

outnumbering males and were located 

off northern NSW, from Newcastle to 

Danger Point, and off Taupo Seamount 

(Graham and Daley, 2011), providing 

some support for spatial sexual 

segregation. Interestingly, Graham and 

Daley (2011) found no evidence of 

sexual or age segregation by depth, with 

males dominating throughout all depth 

zones sampled (with the exception of 

the two sites noted above) and juveniles 

evenly interspersed with adults across 

all depths. 

In terms of mating and reproductive 

behavior, which could provide some 

insight into potential spatial structuring, 

very little information is available. It is 

known that Harrisson’s dogfish is 

viviparous (i.e., gives birth to live 

young), with a yolk-sac placenta. 

Females have litters of one or (more 

commonly) two pups, with size at birth 

around 35–40 cm TL (Graham and 

Daley, 2011). Although the gestation 

period is unknown, a 2 to 3 year period 

has been estimated for other 

Centrophorus species, with continuous 

breeding from maturity to maximum age 

(Kyne and Simpfendorfer, 2007; Graham 

and Daley, 2011). Female C. harrissoni 

mature at sizes around 98 cm TL and 

reach maximum sizes of 112–114 cm 

TL, while males mature around 75–85 

cm TL and reach maximum sizes of 95– 

99 cm TL (Graham and Daley, 2011). 

Female age at maturity is estimated 

between 23 and 36 years of age (Daley 

et al., 2002; Wilson et al., 2009; Last and 

Stevens, 2009; Graham and Daley, 

2011). Longevity is estimated at over 46 

years of age (Wilson et al., 2009). 

Current breeding sites for Harrisson’s 

dogfish are thought to include waters off 

eastern Australia, from Port Stephens to 

31 Canyon, areas off North Flinders and 

Cape Barren in southeastern Australia, 

and waters around Taupo Seamount 

(Williams et al., 2012). These are areas 

where mature males, mature females, 

and juveniles have been recorded, and 

thus are likely to be areas that support 

viable populations where mating and 

pupping occur (Williams et al., 2012). 

However, more extensive sampling, as 

well critical information regarding the 

aspects of the Harrisson’s dogfish 

breeding cycle (including necessary sex 

ratios for successful reproduction, 

preferred mating and breeding grounds, 

and mating and breeding behaviors), is 

needed to identify and fully 

comprehend the spatial dynamics of 

Harrisson’s dogfish. 

For management purposes, 

Harrisson’s dogfish in Australia have 

been separated into two stocks that are 

considered to be ‘‘distinct’’ populations: 

A ‘‘continental slope’’ stock that occurs 

continuously along the Australian 

eastern continental margin, and a 

‘‘seamount stock’’ that occurs on the 

Tasmantid Seamount Chain off NSW 

and southern Queensland, including the 

Fraser, Recorder, Queensland, Britannia, 

Derwent Hunter, Barcoo, and Taupo 

Seamounts. However, to date, no genetic 

studies have been conducted to confirm 

that these two populations are 

genetically distinct, and tagging studies 

are limited, with insufficient recapture 

rates to make any determination 

regarding the connectivity of the 

populations. In addition, there are a 

number of other uncertainties associated 

with the assumption of two separate 

Harrisson’s dogfish stocks, including 

necessary sex ratios and other 

successful reproduction requirements, 

which are further discussed in Miller 

(2014). Due to these uncertainties, we 

do not find conclusive evidence of 

separate populations of Harrisson’s 

dogfish. Therefore, we consider the 

available information for these two 

stocks, including estimates of depletion 

rates and protection benefits of 

management measures, together when 

we determine the status of the entire 

species throughout its range. 

Because species-specific historical 

and current abundance estimates are not 

available, Williams et al. (2013a) used a 

variety of methods and analyses to 

estimate the pre-fishery (pre-1980s) and 

current abundance (in biomass units) at 

fishery stock and sub-regional scales 

(detailed information on the data 

sources and methods can be found in 

Williams et al. (2013a)). Results from 

the various analyses revealed that 

Harrisson’s dogfish is currently 

estimated to be at 21 percent of its prefishery population size throughout its 

core range (with a lower estimate of 11 

percent and upper estimate of 31 

percent). The authors note that this 

overall estimate of decline is strongly 

influenced by the small declines 

estimated on seamounts (Williams et al. 

2013a). The continental margin 

population is estimated to be at 11 

percent of its pre-fishery population size 

(range of 4 to 20 percent; with the 

estimate influenced by uncertainty 

surrounding the level of cumulative 

fishing effort off the northern NSW 

slope). The seamount population is 

estimated to be at 75 percent of its prefishery population size (range 50 

percent to 100 percent). 

Summary of Factors Affecting 

Harrisson’s Dogfish 

Available information regarding 

current, historical, and potential threats 

to Harrisson’s dogfish were thoroughly 

reviewed (Miller, 2014). We find that 

the main threat to the species is 

overutilization for commercial 

purposes, with the species’ natural 

biological vulnerability to 

overexploitation exacerbating the 

severity of the threat, and hence also 

identified as a secondary threat 

contributing to the species’ risk of 

extinction. We summarize information 

regarding these threats and their 

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interactions below, according to the 

factors specified in section 4(a)(1) of the 

ESA. Available information does not 

indicate that habitat destruction, 

modification, or curtailment, disease, or 

predation are operative threats on this 

species; therefore, we do not discuss 

those further here. Because new 

regulatory measures were just recently 

implemented, the adequacy and 

effectiveness of existing regulatory 

measures is discussed in the ‘‘Protective 

Efforts’’ section below. See Miller (2014) 

for full discussion of all threat 

categories. 

Overutilization for Commercial, 

Recreational, Scientific, or Educational 

Purposes 

Historically, Harrisson’s dogfish and 

other gulper sharks were taken in both 

Australian Commonwealth-managed 

commercial trawl fisheries (those that 

are managed by the Australian Federal 

Government, in coordination with 

Australian State fisheries agencies, 

through the Australian Fisheries 

Management Authority (AFMA) (Kyne 

and Simpfendorfer, 2007)) and Statemanaged commercial trawl fisheries 

operating on the upper slope off eastern 

Australia, within the core range of 

Harrisson’s dogfish. Unfortunately, little 

information is available on the specific 

catch of these deep-water sharks, 

primarily due to the historical 

inaccuracy of data reporting and species 

identification issues. These 

Commonwealth and State-managed 

commercial trawl fisheries developed 

off NSW in the 1970s and off Victoria 

and Tasmania in the 1980s. By the early 

1980s, more than 100 trawlers were 

operating off NSW, with around 60 

percent regularly fishing on the upper 

slope. In fact, between 1977 and 1988, 

catches from these upper-slope trawl 

operations comprised more than half of 

the total trawl landings in NSW 

(Graham et al., 2001). Large numbers of 

C. harrissoni were likely caught and 

discarded off NSW during this time, due 

to the absence of a market for deepwater 

shark carcasses (a result of mercury 

content regulations and preference for 

more marketable bony fishes) (Daley et 

al., 2002; Graham and Daley, 2011). 

Similarly, trawlers operating on the 

upper-slope off eastern Victoria reported 

minimal catches of Centrophorus 

dogfishes, but also likely discarded 

substantial numbers due to Victorian 

State restrictions on mercury content in 

shark flesh (Daley et al., 2002). Graham 

and Daley (2011) estimate that landings 

of Centorphorus spp. were around 

several hundred tonnes per year during 

the 1980s and early 1990s. 

Daley et al. (2002) note that in the 

early 1990s significant quantities of 

Centrophorus spp. were also caught off 

eastern Victoria by fishermen using 

droplines targeting blue-eye trevalla 

(Centrolophus antarctica) and ling 

(Genypterus blacodes). In addition, 

some Southern and Eastern Scalefish 

and Shark Fishery (SESSF) operators off 

Victoria used deep-set gillnets to target 

Centrophorus species for their livers in 

the 1990s (Daley et al., 2002). Squalene 

oil, which is extracted from the liver of 

deep-sea sharks, is used in a number of 

cosmetics and health products, and the 

livers of Centrophorus species have the 

highest squalene oil content (67–89 

percent) of any deep-sea shark. 

Fishermen would keep the livers of the 

Centrophorus spp. and discard the 

carcasses due to their mercury content. 

However, by the time the mercury 

restrictions were eased in 1995 

(allowing for carcasses to also be sold), 

very few Centrophorus species were 

being caught off eastern Victoria, with 

targeting of these sharks having 

essentially ceased (Daley et al., 2002). 

Since 2002, total catch of gulper sharks 

by Commonwealth licensed vessels has 

been less than 15 t per year (Woodhams 

et al., 2013). 

In 2001, Graham et al. (2001) 

quantified the effects of the historical 

trawling on the abundance of gulper 

sharks off NSW using data from fisheryindependent surveys conducted along 

the upper slope before and after the 

expansion of the commercial trawlfishery (Andrews et al., 1997). The 

initial pre-fishery survey was carried 

out during 1976 and 1977. There were 

three trawling survey grounds: (1) 

Sydney-Newcastle, (2) UlladullaBatemans Bay, and (3) Eden-Gabo Island 

and eight depth zones (covering depths 

of 200–650 m). The two northern 

grounds (Sydney and Ulladulla) were 

surveyed twice in 1976 and twice in 

1977; the southern (Eden) ground was 

surveyed three times in 1977. These 

surveys were repeated in 1996–1997, 

(with two surveys conducted off Sydney 

and Ulladulla and three off Eden) using 

the same vessel and trawl gear and 

similar sampling protocols, to examine 

the changes in relative abundances of 

the main species (number and kg per 

trawling hour) after 20 years of trawling 

(see Andrew et al., 1997; Graham et al., 

2001). Results from these surveys show 

that Harrisson’s dogfish were present 

and, at one time, were caught across all 

of the survey grounds and depth zones. 

In 1976, catches of Harrisson’s dogfish 

were combined with southern dogfish 

(C. zeehaani) in the initial two surveys 

off Sydney and one off Ulladulla. When 

these species were separated in the later 

1976 surveys, and in 1977, southern 

dogfish comprised around 75 percent 

and Harrisson’s dogfish comprised 25 

percent of the combined catch. In 1976– 

77, Harrisson’s and southern dogfishes 

combined represented around 9 percent, 

18 percent, and 32 percent of the total 

fish catches off Sydney, Ulladulla, and 

Eden, respectively. The overall mean 

catch rate (for all grounds and depths) 

was 126 kg/hour. This is in stark 

contrast to the 0.4 kg/h catch rate in 

1996–1997, when only 14 southern and 

8 Harrisson’s dogfishes were caught, 

comprising 0.18 percent of the total fish 

catch weight (Graham et al., 2001). For 

the 1976–77 surveys where the two 

species were separated, the mean catch 

rate of Harrisson’s dogfish was 28.8 kg/ 

hr caught over the course of 173 tows. 

In 1996–97, the mean catch rate of 

Harrisson’s dogfish was 0.1 kg/hr over 

the course of 165 tows (Graham et al., 

1997; 2001). These decreases in survey 

catch rates provide compelling evidence 

of declines of over 99.7 percent in 

relative abundance of C. harrissoni on 

the upper-slope of NSW, a core part of 

their range, after 20 years of trawling 

activity (Graham et al., 2001). 

In Australia, the commercial trawl 

fisheries are still active, as are demersal 

line fisheries, which also incidentally 

catch Harrisson’s dogfish. In terms of 

Commonwealth-managed fisheries, 

Harrisson’s dogfish are primarily caught 

as bycatch by the SESSF, which 

operates over an extensive area of the 

Australian Fishing Zone (AFZ) around 

eastern, southern, and southwestern 

Australia. The distribution of recent 

(2006–2010) commercial fishing effort 

in the SESSF shows that there is still 

substantial fishing effort on 

Commonwealth upper-slope grounds 

using demersal gears, specifically trawl 

and auto-longline operations (see Miller 

(2014) for more details). According to 

Graham (2013), around 30 trawlers and 

3 auto-longliners in the SESSF still 

operate along the upper-slopes. Since 

auto-longline vessels, which deploy up 

to 15,000 hooks per vessel per day, can 

operate on the steep and rough ground 

that would potentially be a refuge for C. 

harrissoni from trawling (R. Daley, 

Commonwealth Scientific and 

Industrial Research Organization 

(CSIRO), personal communication, 

2014), the combined operation of both 

the trawl and auto-longline fisheries 

within the range of Harrisson’s dogfish 

significantly increases the likelihood of 

incidental catch of the species. Catch 

rates of Harrisson’s dogfish in the 

SESSF have been minimal in recent 

years, likely due to their low abundance 

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on the continental margin; however, the 

combined operation of these demersal 

gears on the upper-slope grounds may 

further decrease abundance of the 

remaining population. For the 2012– 

2013 season, reported gulper shark (C. 

harrissoni, C. moluccensis, C. zeehaani) 

landings (in trunk weight) were 0.9 t 

with discards of 1.2 t (Woodhams et al., 

2013). This is a decrease from the 

previous year, which reported landings 

of 3.8 t. Given the evidence of 

substantial depletion of both Harrisson’s 

and southern dogfishes in Australian 

waters over the years, high risk of 

overfishing in the SESSF, with no 

current indication of recovery (based on 

2012–2013 season data), the Australian 

Government Department of Agriculture 

classified the above three gulper sharks 

as ‘‘overfished’’ in 2012, with the 

current level of fishing mortality noted 

as ‘‘uncertain’’ (Woodhams et al., 2013). 

In fact, upper-slope gulper sharks have 

been classified as overfished since they 

were first included in Australia’s 

Fishery Status Reports in 2005 

(Woodhams et al., 2011). In February 

2013, a zero retention limit was 

implemented for Harrisson’s dogfish 

(Woodhams et al., 2013), along with 

other management measures detailed in 

AFMA’s Upper-Slope Dogfish 

Management Strategy (AFMA, 2012) 

and evaluated in the ‘‘Protective Efforts’’ 

section below. 

In terms of state-managed fisheries, 

the range of Harrisson’s dogfish extends 

within NSW, Victoria, and Tasmania 

jurisdictions. In both Victorian and 

Tasmanian fisheries, catch records of 

Harrisson’s dogfish are rare and 

interactions with these fisheries are 

considered to be unlikely, based on 

their respective fishing operations 

(Threatened Species Scientific 

Committee (TSSC), 2013). In NSW 

commercial fisheries, Harrisson’s 

dogfish may be caught by the Ocean 

Trap and Line Fishery and the Ocean 

Trawl Fishery. According to Graham 

(2013), there are up to five trawlers in 

the Ocean Trawl Fishery that fish 

seasonally between Newcastle and 

Sydney and may incidentally catch 

Harrisson’s dogfish, and only minimal 

line fishing effort on the upper-slope (K. 

Graham, Australian Museum, personal 

communication, 2014). In 2013, a zero 

retention limit was implemented for 

Harrisson’s dogfish (unless for scientific 

purposes as agreed by Fisheries NSW) 

(NSW DPI, 2013). 

Because of their low productivity, 

sustainable harvest rates of gulper 

sharks are estimated to be less than five 

percent of their virgin biomass, and 

maybe even as low as one percent 

(reflecting the proportion of total 

population that can be caught and still 

maintain sustainability of the 

population; Forrest and Walters, 2009). 

However, these harvest levels can only 

be sustained by a population in a 

significantly less depleted state 

(Woodhams et al., 2011). In the case of 

Harrison’s dogfish, Woodhams et al. 

(2013) notes that even low levels of 

mortality can pose a risk because of its 

significantly depleted state. Although 

total fishing mortality on gulper sharks 

is unknown, the level of catch and 

observed discards in recent years was 

deemed likely to result in further 

population declines (Woodhams et al. 

2011; 2012; 2013). In the 2012–13 

fishing season, discards actually 

outnumbered landings (1.2 t compared 

to 0.9 t; Woodhams et al., 2013). Thus, 

even with the prohibition on retention 

of the species, there is still a potential 

for discards based on the significant 

overlap of current fishing effort within 

the core range of the species 

(Woodhams et al., 2013). This is a 

concern because Harrisson’s dogfish 

suffers from high at-vessel mortality in 

trawl gear and potentially high at-vessel 

mortality in auto-longline gear 

(Williams et al., 2013a). Therefore, the 

continued fishing effort on the upperslope and potential for incidental 

capture of Harrisson’s dogfish in the 

trawl and line fisheries described above, 

which will likely result in mortality of 

the species, is considered a threat that 

is currently contributing to the 

overutilization of the species and its risk 

of extinction. 

In the areas off New Zealand where C. 

harrissoni have been observed (Three 

Kings Ridge, Norfolk Ridge, and 

Kermadec Ridge), there is limited 

fishing effort (Graham, 2013). The 

fishing activities include trawling on the 

West Norfolk Ridge, drop-lining for 

large bony fishes on the Three Kings 

Rise, West Norfolk Ridge, and 

Wanganella Bank, and minimal 

longlining and close to no trawling on 

the Kermadec Ridge. No bycatch of 

gulper sharks has been reported from 

these fishing activities (based on a 

personal communication from C. Duffy 

in Graham (2013)). Given the 

uncertainty surrounding the C. 

harrissoni abundance in this area, it is 

currently unknown if these fishing 

activities are impacting Harrisson’s 

dogfish populations or significantly 

contributing to its extinction risk 

(Graham, 2013). 

Other Natural or Manmade Factors 

Affecting the Continued Existence of 

Harrisson’s Dogfish 

Many sharks are biologically 

vulnerable to overexploitation due to 

their life history parameters. Species 

with slow population growth rates, late 

age at maturity, long gestation times, 

low fecundity, and higher longevity are 

especially sensitive to elevated fishing 

mortality (Musick, 1999; Garcı ́a et al., 

2008; Hutchings et al., 2012). These life 

history traits increase the species’ 

susceptibility to depletion by decreasing 

the species’ ability to rapidly recover 

from exploitation. Harrisson’s dogfish 

exhibits these same life history traits, 

with late maturity, long gestation times, 

small litter sizes, and high longevity. 

These life history traits have 

exacerbated the overall impact of the 

historical overutilization of the species 

on its extinction risk, leading to the 

substantial decline in Harrisson’s 

dogfish abundance, and will continue to 

place the species at increased risk of 

demographic stochasticity. 

Extinction Risk 

It is clear that the species faces 

current demographic risks that greatly 

increase its susceptibility to extinction. 

Due to the significant decline, the 

species is no longer found in 

approximately 19 percent of its 

Australian range and, furthermore, 

throughout the rest of its core range, is 

estimated to be at 21 percent of its total 

virgin population size (with separate 

estimates of 11 percent for the 

continental margin population and 75 

percent for the seamount population) 

(Williams et al., 2013a). Although the 

population on the seamounts may be 

less depleted, it also likely comprises a 

significantly smaller portion of the 

entire Harrisson’s dogfish population, 

based on the amount of available habitat 

and corresponding carrying capacity. In 

fact, the continental margin habitat, 

where the population is estimated to be 

at only 11 percent of its total virgin 

population size, represents 86 percent of 

Harrisson’s dogfish’s estimated extent of 

occurrence and 84 percent of its 

estimated area of occupancy (TSSC, 

2013), indicating significant depletion 

throughout most of the species’ range. In 

addition, the existing Harrisson’s 

dogfish populations along the 

continental margin and off the 

seamounts in Australia and New 

Zealand are small and fragmented, with 

only three identified remnant 

populations that are thought to be viable 

(due to presence of mature males, 

females, and/or juveniles within the 

same area). Two of these populations 

are located off the continental margin 

and the third is off Taupo Seamount. It 

is unclear the extent to which these 

populations can help recover 

Harrisson’s dogfish, as breeding 

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population exchange, and general 

movement of individuals is currently 

unknown. Due to their size and 

isolation, these populations may be at 

an increased risk of random genetic drift 

and could experience the fixing of 

recessive detrimental alleles that could 

further contribute to the species’ 

extinction risk (Musick, 2011). In 

addition, the patchy distribution of 

these populations throughout the 

species’ entire range increases 

susceptibility to local extirpations from 

environmental and anthropogenic 

perturbations or catastrophic events. 

Given the apparent spatial structuring of 

the species and dominance of males in 

the sex ratios at many locations, a 

further reduction in the numbers of 

females at any given site may decrease 

reproductive success and prevent 

population replacement. The species 

has extremely low fecundity (2–3 year 

gestation period resulting in 1 to 2 

pups), slow growth rates, and late 

maturity, all of which contribute to a 

long population doubling time. In a 

severely depleted state, these traits may 

contribute to increasing the species’ 

extinction risk, especially if the species 

is still subject to threats that further 

reduce its abundance. Thus, although 

the species’ biological characteristics 

have allowed it to successfully thrive in 

the past, under the current conditions of 

severely fragmented populations and 

low abundance throughout its range, 

questionable population viability, and 

risk of incidental mortality from 

fisheries, the species’ natural life history 

traits are presently threatening its 

continued existence. Specific 

information is lacking on interactions 

among threats. 

Without considering the effectiveness 

of the recently implemented 

management measures in reducing the 

threat of overutilization and improving 

the status of Harrisson’s dogfish in 

Australian waters (discussed in the 

‘‘Protective Efforts’’ section below), 

Miller (2014) concluded that Harrisson’s 

dogfish is presently at a high risk of 

extinction due to threats of 

overutilization exacerbated by its 

natural biological vulnerability to 

depletion, the interaction of which has 

resulted in significant demographic 

risks to the species. We agree with this 

analysis and find that the species is 

presently in danger of extinction 

throughout its range. Below we evaluate 

formalized conservation efforts that 

have yet to be implemented or to show 

effectiveness to determine whether 

these efforts contribute to making listing 

the species as endangered unnecessary. 

We evaluate these conservation efforts 

using the criteria outlined in PECE. 

Protective Efforts 

The EPBC Act, the Australian 

Government’s central piece of 

environmental legislation, applies to 

any group or individual whose actions 

may have a significant impact on a 

‘‘matter of national environmental 

significance.’’ Any proposed action that 

meets this standard must then be 

assessed to determine its environmental 

impact. Species listed as ‘‘vulnerable,’’ 

‘‘endangered,’’ and ‘‘critically 

endangered’’ under the EPBC Act are 

considered to be matters of national 

environmental significance and receive 

these provisions. 

In 2009, Harrisson’s dogfish was 

nominated for listing under the EPBC 

Act. Its status was reviewed by the 

Threatened Species Scientific 

Committee (TSSC), a committee 

established under the EPBC Act to 

advise the Australian Minister for the 

Environment on the amendment and 

updating of lists of threatened species, 

threatened ecological communities, and 

key threatening processes, and with the 

making or adoption of recovery plans 

and threat abatement plans. In 2013, the 

TSSC concluded that Harrisson’s 

dogfish was eligible for listing as 

endangered under the EPBC Act because 

the species had suffered a severe 

reduction in numbers, with a suspected 

population decline of between 74 and 

82 percent (TSSC, 2013). However, the 

TSSC concluded that the species was 

also eligible for listing as a conservation 

dependent species under the EPBC Act 

because it is the ‘‘focus of a plan of 

management [the Strategy] that provides 

for managed actions necessary to stop 

the decline of, and support the recovery 

of, the species so that its chances of long 

term survival in nature are maximized’’ 

(TSSC, 2013). In May 2013, based on the 

TSSC recommendation, the Minister of 

the Environment officially listed 

Harrisson’s dogfish as a conservation 

dependent species under the EPBC Act. 

This listing means that the species is not 

considered a matter of national 

environmental significance in the 

context of the EPBC Act, and, as such, 

Harrisson’s dogfish are exempt from the 

EPBC Act protective provisions. 

In 2012, AFMA published the UpperSlope Dogfish Management Strategy (the 

‘‘Strategy’’; see AFMA, 2012) to satisfy 

the aforementioned management 

requirements for a conservation 

dependent listing of Harrisson’s Dogfish 

and Southern Dogfish under Australia’s 

EPBC Act. The Strategy, which we 

evaluate below according to the 

guidelines in the PECE (68 FR 15100; 

March 28, 2003), includes regulatory 

management measures designed to 

rebuild the Harrisson’s dogfish 

population above a limit reference point 

of 25 percent of its unfished biomass 

(B25). Setting a recovery time frame was 

deemed not feasible until further 

research on the species is completed; 

however, an interim time frame to reach 

this reference point was estimated based 

solely on the biological characteristics 

of the species (three generation times) 

and equal to 85.5 years (SWG, 2012). 

The outcomes and the effectiveness of 

the Strategy are expected to be 

measured on a biennial basis, as 

detailed in AFMA’s ‘‘Upper-Slope 

Dogfish Research and Monitoring 

Workplan.’’ The workplan for the period 

of 2014–2016 (Workplan 1) focuses on 

the development of a cost-effective 

method for measuring baseline relative 

abundance of gulper sharks and 

recovery over time (AFMA, 2014). This 

output will be assessed as part of the 

Research and Monitoring Workplan 

2014–16 review (proposed time frame of 

July 2014-Dec 2016). Once the 

methodology has been developed, the 

next output (Workplan 2) is expected to 

produce baseline relative abundance 

estimates for Southern and Harrisson’s 

dogfish (proposed time frame for output: 

Jan 2017–Dec 2019). Subsequent 

workplans will provide estimates of 

rebuilding over time and will be 

periodically assessed to ensure that the 

actions within the workplans are 

achieving the desired outputs. Hence, it 

appears it will be a number of years 

before the effectiveness of the Strategy 

will be able to be quantified. As 

outlined in the PECE, we must evaluate 

these conservation efforts that have not 

yet demonstrated effectiveness at the 

time of listing to determine whether 

these efforts are likely to be effective at 

reducing or eliminating threats and 

improving the status of Harrisson’s 

dogfish. Below are the regulatory 

measures from the Strategy that have 

already been implemented by AFMA for 

the conservation of the species (under 

the legal authority of section 41A of the 

Australian Fisheries Management Act 

1991 and implemented under ‘‘SESSF 

Fishery Closures Direction No. 1 2013;’’ 

satisfying the first criteria of the PECE) 

and our subsequent evaluation of their 

likely effectiveness at improving the 

status of Harrisson’s dogfish (the second 

criteria of the PECE). The figures and 

tables referenced below can be found in 

the PECE supplement (Miller, 2014b). 

Prohibition on the Commercial 

Retention of Gulper Sharks 

The Strategy implements a complete 

prohibition on the commercial retention 

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of all gulper sharks. However, even 

before the prohibition, reported catch 

rates of Harrisson’s dogfish in the 

SESSF have been minimal in recent 

years, likely due to the low abundance 

of the species on the continental margin 

where the fisheries operate. Harrisson’s 

dogfish are not a targeted species, but 

rather taken as incidental catch. 

Although this prohibition will decrease 

the numbers of sharks being landed, it 

is worth noting that discards have 

outnumbered landings in recent years 

and at a rate that was deemed likely to 

result in further declines of the species 

(Woodhams et al., 2011). Additionally, 

in the latest Fishery Status Report for 

Commonwealth-managed fish stocks, it 

states: ‘‘[t]here is potential for 

unreported or underestimated discards 

(based on the large degree of overlap of 

current fishing effort with the core range 

of the species [Harrisson’s dogfish]), and 

low levels of mortality can pose a risk 

for such depleted populations’’ 

(Woodhams et al., 2013). Based on the 

above discarding trends, the fact that it 

is the Commonwealth Trawl Sector of 

the SESSF which is the main fishery 

operating within the species’ core 

continental margin range, and the 

evidence that Harrisson’s dogfish are 

not expected to survive after incidental 

capture in trawl gear (Rowling et al., 

2010), the new retention prohibition 

may only have a minor impact on 

decreasing current fisheries-related 

mortality. 

Network of Spatial/Area Closures 

Prior to the Strategy, a number of 

closures were implemented across the 

SESSF operational area (AFMA, 2012); 

however, there were concerns that these 

closures were too small in relation to 

the historical distribution of the species 

to prevent further declines or recover 

the species (Musick, 2011; Woodhams et 

al., 2011). Musick (2011) estimated that 

the closures protected Harrisson’s 

dogfish from all forms of industrial 

fishing in only 9.8 percent of its habitat. 

In response to these concerns, AFMA 

evaluated options for closures in the 

Strategy and created a new network of 

spatial/area closures in 2013, taking into 

account the species’ distribution and 

habitat potential, which would protect 

the species from various forms of fishing 

and prevent further declines. 

Regulations that are the most effective 

in protecting the species from threats of 

overutilization (i.e., incidental catch) 

are those that prohibit all types of 

fishing methods. An analysis of already 

implemented conservation efforts from 

the Strategy estimates that 26.3 percent 

of the core Harrisson’s dogfish seamount 

habitat (weighted by carrying capacity— 

the habitat area’s ability to support 

dogfish populations) and 5.5 percent of 

the continental margin habitat are 

closed to all types of fishing (see Table 

1; Figures 1 and 4 in Miller, 2014b). In 

terms of the areas that support 

Harrisson’s dogfish populations, this 

coverage translates to protection for 26.3 

percent of the current biomass of the 

seamount population (provided by the 

new Derwent Hunter closure) and 19.1 

percent of the biomass of the 

continental margin population. 

Contributing to the protection of the 

continental margin population are the 

Strategy’s extension of the Flinders 

Research Zone closure and revision to 

the Harrisson’s Gulper closure that 

prohibits fishing in the depth range of 

Harrisson’s dogfish. The fact that these 

closures encompass areas critical to 

population viability further increases 

the effectiveness of this regulation in 

improving the status of the species. For 

example, the Extended Flinders 

Research Zone (see Figures 2a and 2b in 

PECE supplement) protects the only 

known potentially reproducing 

population of Harrisson’s dogfish found 

south of Sydney. Specifically, this 

closure protects the mature male 

population found around Babel Island, 

the mature female population found 

around Cape Barren, and the likely 

migration route between these two 

populations that is thought to support 

mating activities (Middle Ground). Prior 

to this closure, only the Babel and Cape 

Barren grounds were protected, leaving 

the closely adjacent Trawl Corridor and 

Middle Ground open to fishing 

activities (and the potential for 

incidental catch). Now, this closure has 

been extended and prohibits all fishing 

methods from 200 to 1000 m deep, 

covering the entire depth range of 

Harrisson’s dogfish. 

If we also consider closures that 

prohibit all high-risk fishing methods 

(permitting only power hand-line), the 

protection coverage increases to 24 

percent of Harrisson’s dogfish’s entire 

core habitat (see Table 1; Figures 1–4 in 

Miller, 2014b). The effectiveness of 

these regulations in improving the 

status of Harrisson’s dogfish partly 

depends on the handling of the species 

in fishing gear and subsequent postrelease mortality rates of the shark. In 

other words, these regulations are only 

likely to be effective in decreasing 

threats if they reduce incidental catch 

altogether or reduce mortality rates of 

Harrisson’s dogfish when incidentally 

caught. As these closures prohibit all 

fishing with the exception of powerhandline methods, we need to consider 

the selectivity and post-release mortality 

of power-handline methods on 

Harrisson’s dogfish in order to evaluate 

the effectiveness of these closures. 

Based on findings from Graham (2011) 

and Williams et al. (2013b), there is a 

high selectivity rate for target species 

(and consequently low bycatch) when 

using the power handline technique. 

For example, in one of the experiments 

designed to replicate normal powerhandline fishing operations for 

harvesting blue-eye trevalla (the target 

species for power-handline fishing), 

results showed that Harrisson’s dogfish 

could be successfully avoided. Out of a 

total of 1,435 individual line drops, 

25,509 hooks, and over 10 fishing trips, 

no Harrisson’s dogfish were taken as 

bycatch. This is in contrast to the 6,819 

blue-eye trevalla that were caught using 

the power-handline method (Williams 

et al., 2013b). Likely contributing to this 

high degree of selectivity using the 

power handline method and avoidance 

of Harrisson’s dogfish is the fact that 

fishing for blue-eye trevalla is normally 

conducted during daylight hours, in 

depths of 280–550 m. Based on 

Harrisson’s dogfish’s diel-migration 

patterns, the species is normally found 

in depths greater than 550 m during 

daylight hours, deeper than the normal 

power handline operating depths. 

Insight into post-release mortality was 

also provided from the Williams et al. 

(2013b) study, as exploratory fishing for 

Harrisson’s dogfish was conducted to 

determine the occurrence of the species 

on the seamounts. A total of 105 

Harrisson’s dogfish were captured 

during this exploratory component of 

the survey and Williams et al. (2013b) 

observed that many of these sharks, 

when brought to the surface, were in 

good physical condition. All but one 

shark were released back into the water 

alive and actively swam away. Williams 

et al. (2013b) attribute this potentially 

low post-release mortality to the short 

soak times associated with powerhandline fishing. In addition, this type 

of fishing method consists of a high 

degree of spatial targeting and small 

gear size, which also likely contribute to 

a high survival rate of Harrisson’s 

dogfish when caught on lines (Williams 

et al., 2013b). Based on these findings, 

we consider closures that prohibit all 

high-risk fishing methods (permitting 

only power hand-line), as effectively 

decreasing the threat of overutilization 

(i.e., mortality from incidental catch) of 

Harrisson’s dogfish (see Table 1; Figures 

1–4 in Miller, 2014b). The coverage of 

these closures, when broken out by 

continental margin and seamount 

proportions and weighted by carrying 

capacity, translates to protection for 

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Harrisson’s dogfish over 18.4 percent of 

its core continental margin habitat and 

77.6 percent of its seamount habitat (see 

Table 1 in Miller, 2014b). Contributing 

to the protection of the continental 

margin population is the Strategy’s 

extension of the Endeavour closure, and 

for the seamount population, the newly 

created Queensland and Britannia 

seamount closures. 

If we look at the closures that prohibit 

trawling operations next, it is estimated 

that 29.5 percent of the species’ core 

habitat range is protected from trawling 

activities (see Table 1 in Miller, 2014b). 

With these regulations, almost all of the 

Harrisson’s dogfish’s core seamount 

habitat would be protected. As 

Harrisson’s dogfish are not expected to 

survive when caught in trawl gear, these 

closures are likely to be effective in 

decreasing mortality rates from 

incidental catch in trawls. In fact, there 

is already evidence of rebuilding in 

areas that were extensively trawled but 

have seen significantly less activity 

recently. Graham and Daley (2011) note 

the presence of a high numbers of 

juveniles (<80 cm TL, including 

neonates) that were caught during a 

2009 long-line survey at sites off Port 

Stephens NSW. This area had been 

extensively trawled during the first 20 

years of the upper-slope fishery, but 

over the last 10 years has seen 

significantly less trawling activity 

(Graham et al., 2001; Graham and Daley, 

2011). The authors of the study attribute 

the increase in juvenile sightings as 

potentially a re-establishment of the 

population in this area. 

NSW closures and regulations may 

also offer additional protection to the 

species (TSSC, 2013). Specifically, the 

NSW ‘‘North of Sydney closure’’ (see 

Figure 3 in Miller, 2014b) prohibits all 

fishing methods except for powerhandline, but allows trawling in depths 

over 650 m (which overlaps with the 

Harrisson’s dogfish depth range). The 

NSW trawl restriction areas 4 and 6 (see 

Figure 5 in Miller, 2014b) also provide 

some protection by prohibiting trawling, 

but are open to line methods. Overall, 

these additional regulations protect 2.4 

percent of the core habitat (and 3 

percent of the core continental margin 

habitat), mainly from trawling, except at 

the shallowest depths (TSSC, 2013). 

Many uncertainties surround these 

estimates. We currently do not know the 

locations of important foraging grounds 

or nursery areas that are critical for 

population viability. In addition, we 

have no information regarding the 

movement of Harrisson’s dogfish in and 

out of these protective closures, or the 

connectivity between the seamounts 

and continental margin populations. 

However, preliminary tagging studies of 

a closely related species, C. zeehaani, 

inside a fishery closure off southern 

Australia suggest that the home ranges 

of deep-water dogfish sharks may be 

small, with evidence of resident female 

populations that can be effectively 

protected by fishery closures (Daley et 

al., 2014). Furthermore, as new 

information becomes available that 

improves the understanding of 

Harrisson’s dogfish biology and stock 

structure, the management arrangements 

in the Strategy can be adapted as 

necessary to ensure the effectiveness of 

the Strategy over time. 

Compliance and Enforcement 

In addition to the actual spatial extent 

of the closure network, the certainty of 

effectiveness of these regulatory 

measures in decreasing threats to the 

species also depends on the compliance 

and enforcement of these closures. For 

the Commonwealth fisheries, AFMA has 

created a compliance team to assist with 

issues such as quota evasion and 

balancing, Vessel Monitoring System 

(VMS) requirements, and compliance 

with fisheries closures and interactions 

with protected species. In terms of VMS 

requirements (a key monitoring 

provision in the Strategy), compliance 

rates have significantly increased over 

the years, thanks to outreach material to 

vessel operators. Compliance rates for 

the requirement for vessels to have an 

operational VMS averaged around 97 

percent for the 2012–2013 year (AFMA, 

2013a). 

Another key to the successful and 

effective conservation of the Harrisson’s 

dogfish population so that it may 

rebuild in the future is compliance with 

fishing prohibitions inside closures. In 

2010–2011, AFMA identified the 

activity of fishing boats entering and/or 

fishing inside closures as an occasional 

but significant risk. To combat this, they 

developed a ‘‘show cause’’ program 

whereby breaches inside closures were 

identified from VMS, and the operators 

of these vessels were sent a letter asking 

them to explain or ‘‘show cause’’ for 

their activity. Within a year of running 

the program, the incidence of fishing or 

navigating inside fishery closures had 

decreased from an average of 11 

breaches per month to less than 2 

breaches per month (AFMA, 2013b). 

Conclusion 

After consideration of the evaluation 

criteria for certainty of effectiveness 

under the PECE, we find that these 

existing regulatory measures are likely 

to be effective in improving the present 

status of the species. The network of 

implemented closures addresses the 

threat of overutilization by prohibiting 

high-risk fishing methods, which 

decreases fishery-related mortality from 

bycatch. Based on a prior review by 

Musick (2011), it was recommended 

that closures include at least 20 to 35 

percent of important Harrisson’s dogfish 

habitat in order to prevent further 

decline of the species and potentially 

support recovery. Overall, the closures 

evaluated above appear to provide the 

species with effective protection from 

high-risk fishing methods over 24 

percent of its core habitat range (see 

Table 1 in Miller, 2014b). Specifically, 

the core habitat of the much-lessdepleted seamount population is 

significantly protected from high-risk 

fishing methods and almost entirely 

protected (98.2 percent) from trawling 

activities (see Table 1 in Miller, 2014b). 

In fact, 77.6 percent of the seamount 

population biomass is protected from all 

high-risk fishing methods by the new 

closures created by the Strategy. These 

conservation efforts are likely to 

effectively improve and protect the 

status of this population so that it is no 

longer presently in danger of extinction. 

In terms of the continental margin 

population, the new network of spatial 

closures provides protection from highrisk fishing methods over 18.4 percent 

of the core margin habitat. The closures 

protect 32.4 percent of the current 

biomass, including the only known 

viable population found south of 

Sydney, from all fishing activities, 

which will be critical for improving the 

status of the population (see Table 1; 

Figure 1 in Miller, 2014b). Although 

incidental fishing mortality may occur 

outside of these closures, based on the 

best available information, we consider 

the current network of closures effective 

in adequately decreasing the present 

threat of overutilization throughout the 

species’ range to the point where the 

species is not currently in danger of 

extinction. 

As mentioned previously, these 

conservation efforts have been designed 

with the explicit objective to stop the 

decline of Harrisson’s dogfish and 

rebuild the population above 25 percent 

of its unfished biomass. AFMA’s 

‘‘Upper-Slope Dogfish Research and 

Monitoring Workplan’’ details the 

provisions for monitoring and reporting 

progress on the objective and 

effectiveness (based on evaluation of 

quantifiable parameters and using 

principles of adaptive management) of 

the implemented conservation efforts. 

Specifically, the outcomes and the 

effectiveness of the Strategy are 

expected to be measured on a biennial 

basis. However, as noted below, 

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certainty that the above conservation 

efforts will remain in place after 5 years 

cannot be predicted at this time. As it 

stands, the Strategy, and conservation 

efforts therein, are only a force under 

Australian law if AFMA continues to 

implement the closures under section 

41A of the Fisheries Management Act 

1991. These closures are implemented 

under ‘‘Directions’’ (for example, the 

current fishery closures to protect 

Harrisson’s dogfish have been 

implemented under ‘‘SESSF Fishery 

Closures Direction No. 1 2013’’). These 

legal instruments are only in effect for 

5 years, after which AFMA may choose 

to extend the closures by creating a new 

Direction. If AFMA does not take action 

after 5 years, these closures will expire. 

Although the Upper-Slope Dogfish 

Research and Monitoring Workplan 

details AFMA’s commitment to stop the 

decline of Harrisson’s dogfish and work 

to rebuild the population, the protection 

of the species is not required under the 

EPBC Act since the species was listed as 

conservation dependent instead of 

endangered. In addition, in the case 

where any part of this Strategy ceases to 

exist or changes, the species would not 

automatically be listed as endangered 

under the EPBC Act. Rather, the TSSC 

would be convened and asked to 

evaluate how the changes impact the 

status of the species and provide 

recommendations on listing eligibility 

to the Minister for the Environment, 

with the ultimate decision on whether 

to list the species in a given category 

made by the Minister. 

While we conclude that the present 

conservation efforts are currently 

effective in preventing the extinction of 

the species, we have no certainty that 

they will remain in place after 5 years. 

Taking into account the present state 

and life history of the species, we do not 

consider 5 years to be sufficient time for 

the status of the species to improve to 

where it is no longer in danger of 

extinction without the continued 

implementation of these efforts. In other 

words, the removal of these 

conservation efforts after 5 years will 

once again subject the species to the 

threats described previously, and based 

on the information from the extinction 

risk analysis (e.g., substantial depletion, 

fragmented populations, extremely low 

productivity, sensitivity to low levels of 

mortality), we find that the species will 

likely become in danger of extinction at 

that time. 

In conclusion, after consideration of 

the evaluation criteria under the PECE, 

we are sufficiently certain that the 

implemented conservation efforts will 

effectively decrease the threat of 

overutilization by fisheries in the near 

term to the point where the species is 

no longer presently in danger of 

extinction. However, given that the 

implementation of these conservation 

efforts is only certain for 5 years, a time 

frame that is insufficient to increase the 

species’ chances of survival when faced 

again with prior threats, we conclude 

that the species will likely be in danger 

of extinction in the foreseeable future. 

We specifically seek additional 

information from the public comment 

process on these conservation efforts 

and their certainty of implementation 

and effectiveness (see below). 

Proposed Determination 

We assessed the ESA section 4(a)(1) 

factors and conclude that the species 

faces ongoing threats from 

overutilization, with the species’ natural 

biological vulnerability to 

overexploitation exacerbating the 

severity of the threats. The species faces 

demographic risks, such as small and 

fragmented populations with low 

productivity, which make it likely to be 

influenced by stochastic or depensatory 

processes throughout its range and place 

the species in danger of extinction from 

the aforementioned threats. We deem 

ongoing conservation efforts as 

currently effective in decreasing the 

main threat of overutilization to the 

point where the species is no longer 

presently in danger of extinction. 

However, the time frame over which 

these conservation efforts will certainly 

be in place is insufficient to increase the 

species’ chances of survival or prevent 

its extinction through the foreseeable 

future. Therefore, based on the best 

available scientific and commercial 

information as presented in the status 

report and this finding, we find that C. 

harrissoni is not currently in danger of 

extinction throughout its range, but is 

likely to become so in the foreseeable 

future. We propose to list Harrisson’s 

dogfish as a threatened species under 

the ESA. 

Corals 

The three coral species considered 

herein are all marine invertebrates in 

the phylum Cnidaria. The phylum is 

called Cnidaria because member species 

use cnidae (capsules containing stinging 

nematocysts) for prey capture and 

defense. All are tropical, shallow water, 

scleractinian (‘‘stony’’) corals that 

secrete a calcium carbonate skeleton. 

Two of the three have the typical stony 

coral symbiosis with zooxanthellae 

(photosynthetic) algae that reside in 

gastrodermal cells of the coral tissue. 

All are non-reef building corals that live 

in small colonies or as solitary 

individuals. The following section 

describes our analysis of the status of 

the three species. Information on many 

of the species is sparse, so we cannot 

provide complete descriptions of their 

natural history. More details can be 

found in Meadows (2014). 

Species Description of Cantharellus 

noumeae 

Cantharellus noumeae is a fungiid or 

mushroom coral that was the first 

described species of its genus, in 1984 

(Hoeksema and Best, 1984). It received 

its own new genus name because, 

unlike most other fungiid corals, it is 

stalked and not free-living as an adult. 

Other species in the genus have since 

been discovered and named, so the 

genus is no longer monotypic. Polyps 

are relatively small for a fungiid coral, 

ranging from 25 to 65 mm in diameter 

(Hoeksema and Best, 1984). The polyps 

are cup-shaped when fully developed 

and have wavy margins (AIMS, 2013a). 

The primary septa are thin. The species 

may be solitary or colonial; colonies 

consist of a few contorted polyps. Their 

typical color is mottled brown. 

Cantharellus noumeae was thought to 

occur only in a restricted area of less 

than 225 km 2 on reefs in sheltered bays 

in New Caledonia, on the southern tip 

of the main island of Grand Terre 

(Hoeksema et al., 2008). Recent research 

by the French Institut de Recherche 

pour le De ́veloppement (IRD) has found 

that the species also occurs on fringing 

reefs farther up the southeast coast at 

Noumea and at Balabaio in the 

northeastern part of New Caledonia 

(www.lagplon.ird.nc; Antoine Gilbert, 

Ginger Soproner, personal 

communication, 2013). It is found in 

waters 10 to 35 m deep, close to soft 

sediment habitats that are in sheltered 

bays and lagoons (Hoeksema and Best, 

1984). There are records of it in western, 

northern, and eastern parts of the island 

of New Guinea that includes Papua New 

Guinea and West Papua, Indonesia, with 

details likely to be published soon on a 

new Web site (http://

coralsoftheworld.com; Charlie Veron, 

personal communication). There are 

also reports of it from Papua New 

Guinea in the International Union for 

Conservation of Nature (IUCN) 

assessment, but the assessment 

questions the validity of this record 

(Hoeksema et al., 2008). The IUCN 

assessment and the researcher whose 

published record is in question (Doug 

Fenner) suggest further confirmation is 

necessary (Hoeksema et al., 2008; 

Fenner, personal communication). 

Fossil records from over 5 million years 

ago indicate that this species was at one 

time found as far west as East 

Kalimantan, on the island of Borneo, 

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Indonesia (Hoeksema, 1989; Hoeksema, 

1993). 

Scleractinian corals have diverse 

reproductive strategies, including both 

asexual and sexual modes of 

reproduction (see Brainard et al., 2011). 

Individual reproductive modes for these 

three species have not been studied. 

Cantharellus noumeae may be a 

sequential sex-changing species like 

other members of its family. Because of 

their relationship with symbiotic 

zooxanthellae, C. noumeae needs to live 

in shallow water to be exposed to light 

the symbiotic algae use to 

photsynthetically fix carbon. 

There is no quantitative speciesspecific population or trend information 

available for C. noumeae (Hoeksema et 

al., 2008; Gilbert, personal 

communication). The current and 

continuing presence of the species in 

New Caledonia was confirmed by Bert 

Hoeksema (personal communication) in 

2012 and in one murky location in 

Prony Bay on the southern tip of Grand 

Terre in 2013 (Andrew Bruckner, 

personal communication). In addition, 

Antoine Gilbert (personal 

communication) notes that from surveys 

he has done over the past 4 years, the 

species is ‘‘uncommon and usually 

found in fringing reefs where 

sedimentation is quite intense.’’ He also 

noted that the species is ‘‘usually found 

in low density, [but] it was observed in 

relative[ly] high density on the slope of 

artificial shores (embankment) in the 

biggest (commercial and industrial) 

harbour of New Caledonia: la Grande 

Rade.’’ We found no information on 

abundance or trends on New Guinea. Its 

presence at one site in Milne Bay 

(Fenner, 2003) is uncertain; Charlie 

Veron may publish information from 

New Guinea on his Web site soon (see 

above). 

Species Description of Siderastrea 

glynni 

Siderastrea glynni was described in 

1994 (Budd and Guzma ́n, 1994). It 

occurs in non-reef-forming spherical 

colonies that are 70 to 100 mm in 

diameter (AIMS, 2013b). They have 

polygonal corallites that are 2.5 to 3.5 

mm in diameter (Budd and Guzma ́n, 

1994). The species is a light reddishbrown in color and occurs on coarse 

sand-rubble substrates. Recent genetic 

work by Forsman et al. (2005) has 

shown that S. glynni is genetically very 

similar to the Caribbean species S. 

siderea, though there are differences 

between the species. Their study could 

not differentiate between two possible 

explanations of the species’ evolution: 

(1) that S. siderea and S. glynni are the 

same species and that S. glynni may 

have recently passed through or been 

carried across the Panama Canal to the 

Pacific Ocean side; or (2) the alternate 

possibility that S. glynni evolved from 

S. siderea, likely about 2 to 2.3 million 

years ago during a period of high sea 

level, when the Isthmus of Panama may 

have been breached, allowing interbasin transfer of the species’ ancestors. 

Because the available information to 

reclassify the species is inconclusive, 

we determine that S. glynni is a valid 

and unique species. 

The range of S. glynni is a small area 

of the Pacific Ocean near the small 

island of Uraba in Panama Bay, a few 

kilometers from the opening of the 

Panama Canal (Guzma ́n and Edgar, 

2008). Identified colonies of S. glynni 

were reported to be unattached and 

occur ‘‘along the upper sand-coral 

rubble reef slope at a depth of 7 to 8.5 

meters’’ (Budd and Guzma ́n, 1994). All 

the islands around the site, as well as 

another set of islands to the south, were 

searched several times without finding 

any additional colonies (Fenner, 2001). 

The reproductive mode for this 

species has also not been studied. 

Because of their relationship with 

symbiotic zooxanthellae, S. glynni need 

to live in shallow water to be exposed 

to light the symbiotic algae use to 

photsynthetically fix carbon. 

Only five colonies of S. glynni have 

ever been found. All were found by 

Budd and Guzma ́n (1994) when they 

discovered the species in 1992. All five 

colonies occurred within a small area of 

less than 10 m 2, with each colony 

within 1 m of another (Budd and 

Guzma ́n, 1994). Each colony was no 

more than 20 cm 2 in size. One colony 

was sacrificed in order to provide 

material for the species’ description. 

During the 1997–98 El Nin ̃o event, the 

four surviving colonies started to 

deteriorate, displaying signs of 

bleaching and tissue loss. Due to their 

unhealthy state, the four colonies were 

moved to Smithsonian Tropical 

Research Institute (STRI) aquaria in 

Panama City, Panama, where they 

remain to this day (Guzma ́n and Edgar, 

2008; Hector Guzma ́n, STRI, personal 

communication, 2013). According to 

Guzma ́n (personal communication, 

2013) the colonies were fragmented to 

increase the number of specimens, but 

their growth rate has been very slow, 

and some fragments did not survive. 

From the original colonies, only one 

survives, with less than 4 cm2 of living 

tissue. Nine of the fragmented colonies 

also survive in the lab and all are less 

than 9 cm 2 in area (Guzma ́n, personal 

communication, 2013). No known 

colonies exist in the wild; however, 

there is a possibility that it still exists 

elsewhere in the wild and is yet 

undiscovered (Guzma ́n and Edgar, 

2008). There are no plans to reintroduce the species, as existing 

colonies are too small to survive, though 

three of the fragments are being 

considered for cryopreservation, which 

would further reduce the population 

size (Guzma ́n, personal communication, 

2013). 

Species Description of Tubastraea 

floreana 

Tubastraea floreana was first 

described by Wells (1982). It is an 

azooxanthellate species, which means it 

lacks the symbiotic photosynthetic 

zooxanthellae that most scleractinians 

have. It has a bright pink color while 

alive, but turns deep red-black when 

dead out of water. Corallites in the 

species are closely spaced (Cairns, 1991) 

and about 4–6 mm in size (Wells, 1983). 

Tubastraea floreana is endemic to a 

few sites on a number of islands in the 

Galapagos Islands chain. It is mostly 

found in cryptic habitats, including on 

the ceilings of caves, and on ledges and 

rock overhangs (Hickman et al., 2007). 

It has been reported to occur at depths 

of 2 to 46 m (Hickman et al., 2007). 

The reproductive mode of this species 

has not been studied, but other 

Tubastraea species reproduce asexually. 

Other Tubastraea species are invasive 

and productive (Riul et al., 2013), so T. 

floreana is also likely to be moderately 

productive. 

According to Hickman et al. (2007), 

prior to the 1982–83 El Nin ̃ o Southern 

Oscillation (ENSO) this species was 

known from six sites on four islands in 

the Galapagos. Since the 1982–83 

ENSO, specimens have only been 

observed at two sites. At one of these 

two sites, the species has not been seen 

since 2001, leaving only a single 

confirmed site with living specimens 

(Hickman et al., 2007). Recent reports 

indicate the species is still present in at 

least one site (Stuart Banks, Charles 

Darwin Foundation, personal 

communication, 2013). We know of no 

other published information on 

distribution or abundance for this 

species. 

Summary of Factors Affecting the Three 

Species of Coral 

Next we consider whether any one or 

a combination of the threat factors 

specified in section 4(a)(1) of the ESA 

are contributing to the extinction risk of 

these three corals. Available information 

does not indicate that overutilization is 

an operative threat for these species; 

therefore, we do not discuss this factor 

further here. We discuss each of the 

remaining four factors and their 

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interaction in turn below, with speciesspecific information following a general 

discussion relevant to all of the species. 

A full review of all of the ESA section 

4(a)(1) threat factors can be found in 

Meadows (2014b) and our final rule 

listing 20 corals (20-coral listing rule) 

under the ESA (79 FR 53851; September 

10, 2014), which provides a general 

global summary of threats facing corals. 

Our 20-coral listing rule identified 

ocean warming, ocean acidification, sealevel rise, disease, sedimentation, 

nutrient enrichment, and fishing as the 

major global threats to coral reefs. The 

information about these threats and the 

species’ responses to these threats is 

described in the 20-coral listing rule and 

incorporated herein by reference. 

Species-specific information regarding 

applicability of these threats to the three 

species considered here is discussed 

below, where available. The extent to 

which the risks discussed in the 20- 

coral listing rule are similar to the risks 

to these three corals is discussed for 

each species. 

The Present or Threatened Destruction, 

Modification, or Curtailment of Its 

Habitat or Range 

Habitat modification from climate 

change is a potential threat to all three 

species of corals (79 FR 53851; 

September 10, 2014). Coral bleaching 

occurs when the photosynthetic 

zooxanthellae symbionts of corals are 

damaged by light at higher than normal 

temperatures. The resulting damage 

leads to the expulsion of these 

important organisms from the coral 

host, depriving the host of the nutrients 

and energy provided by the 

zooxanthellae. While corals can survive 

mild to moderate bleaching, repeated, 

severe, or prolonged bleaching can lead 

to colony mortality. Bleaching events 

have been increasing both in intensity 

and geographic extent due to worldwide 

anthropogenic climate change (HoeghGuldberg, 2006; Eakin et al., 2009). 

Certain genera and growth forms, 

particularly branched species, are more 

sensitive to bleaching than others 

(Wooldridge, 2013). Many corals are 

physiologically optimized to their local 

long-term seasonal variations in 

temperatures and an increase of only 1– 

2 °C above the normal local seasonal 

maximum can induce bleaching 

(Brainard et al., 2011; Logan et al., 

2013). The United States NOAA Coral 

Reef Watch satellite bleaching database 

shows that the range of all three species 

occurs in areas that frequently have 

bleaching alerts, with alerts being more 

frequent and severe in the ranges of S. 

glynni and T. floreana, than in the range 

of C. noumeae. 

Ocean acidification threatens to slow 

or halt coral growth and reef building 

entirely if the pH of the ocean becomes 

too low for corals to form their calcite 

skeletons, but tolerance appears to vary 

by species for those that have been 

studied (see Brainard et al., 2011). In 

addition, bioerosion of reefs is likely to 

accelerate as coral skeletons become 

more fragile as a result of the effects of 

acidification, but effects are highly 

species-specific. Since the petitioned 

species are not reef-building, this effect 

is likely to be less significant. 

Sea-level is also likely to rise as a 

result of climate change, but effects on 

corals are highly uncertain, owing to 

uncertainty in both the likely rate and 

extent of sea-level rise as well as the 

ability of corals generally (or the 

petitioned species specifically) to keep 

pace with the rise in sea level (Brainard 

et al., 2011; 79 FR 53851; September 10, 

2014). 

While climate change effects are 

likely to be serious for many corals, 

Brainard et al. (2011) and our final rule 

listing 20 corals under the ESA (79 FR 

53851; September 10, 2014) show that 

adaptation and acclimatization of corals 

to increased ocean temperatures are 

possible, that there is intra-genus and 

inter-species variation in susceptibility 

to bleaching, ocean acidification, and 

sedimentation, that at least some species 

have already expanded their range in 

response to climate change, and that not 

all species are seriously affected by 

ocean acidification. In addition, a more 

recent paper by Logan et al. (2013) 

examined the potential for coral 

adaptation and acclimatization to 

climate change and found that these 

processes can reduce the frequency of 

mass bleaching events in the future. 

Their modeling results suggest some 

adaptation or acclimatization may even 

have already occurred. A study by 

Wooldridge (2014) provides support 

that a suite of morphological and 

physiological traits relate to bleaching 

vulnerability. These include symbionts’ 

type, metabolic rate, colony tissue 

thickness, skeletal growth form, mucus 

production rates, fluorescent pigment 

concentrations, and heterotrophic 

feeding capacity. According to 

Wooldridge (2014), these traits tend to 

correlate with the ends of the dichotomy 

of branching and plate corals with thin 

tissue layers versus massive and 

encrusting corals with thick tissue 

layers. The species under consideration 

here are not necessarily the most 

vulnerable, based on those traits (see 

below). Therefore, while climate change 

is generally considered a potential 

threat to these candidate corals, the 

likelihood and magnitude of threats 

from climate change are largely speciesspecific and must be examined on that 

basis to fully assess extinction risk (79 

FR 53851; September 10, 2014). 

In addition to the general global 

threats identified in our status review of 

82 corals and final rule listing 20 corals 

under the ESA (Brainard et al., 2011; 79 

FR 53851; September 10, 2014), there 

are some species-specific threats for 

which we have detailed information at 

the scale of these species’ ranges that are 

discussed below. 

Cantharellus noumeae 

Cantharellus noumeae is exposed to 

deforestation, urbanization, and mining 

activity that causes sedimentation and 

water pollution throughout its range in 

New Caledonia (Hoeksema et al., 2008; 

David et al., 2010; McKenna et al., 

2011). The mining activity is a result of 

nickel and smaller amounts of other 

metal mining (cobalt and chromium 

especially) on land throughout the main 

island of Grand Terre (McKenna et al., 

2011; Hoeksema, personal 

communication). Grand Terre holds 25 

percent of the world’s known nickel 

deposits (McKenna et al., 2011). Nickel 

mining started there in the 1870s. 

Currently, most mining is done by opencast strip mining, which has caused 

deforestation and increased erosion and 

runoff of sediments leading to varying 

degrees of sedimentation and light 

attenuation throughout the lagoon of 

Grand Terre, including in areas in and 

adjacent to the species’ range (Ouillon et 

al., 2010). Labrosse et al. (2000) estimate 

that 300 million m 3 of soil has been 

displaced since the beginning of mining 

activities. Mines are located across the 

country, including the large new Goro 

complex, which includes mines, 

processing facilities, and a port. The 

complex began production in late 2010 

and is very near the most abundant 

population of C. noumeae. The Goro 

complex has already had three incidents 

affecting the environment, involving 

spills or releases of sulfuric acid 

solutions used in the processing of the 

nickel ore (Sulfuric Acid on the Web, 

2013). Runoff of heavy metals from the 

mining operations has greatly increased 

concentrations of those metals in the 

marine environment (Fichez et al., 

2010). Nickel has been shown to affect 

fertilization success of four reef coral 

species in the families Acroporidae and 

Faviidae (Reichelt-Brushett and 

Harrison, 2005) and to affect settlement 

and cause mortality of larvae in the 

coral Pocillopora damicornis (Goh, 

1991). Gilbert (personal communication, 

2013) reports that the species is 

common in areas of high sedimentation 

and in the largest harbor, so it may be 

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tolerant to environmental stressors like 

sedimentation. The species may have 

the ability to actively remove sediments, 

as has been shown in some other 

fungiid corals (Bongaerts et al., 2012), 

but this is uncertain. Mitigation 

measures for mining operations are 

required by legislation and include reef 

monitoring requirements (UNESCO, 

2011; Gilbert, personal communication, 

2013), but this monitoring is not at the 

species level (Gilbert, personal 

communication, 2013). It is unclear how 

effective the mitigation methods are, as 

sedimentation and pollution remain 

concerns (David et al., 2010). 

Despite the frequency of bleaching 

alerts, heat-related bleaching is 

apparently not a significant current 

threat in the range of C. noumeae in 

New Caledonia, as water temperatures 

there are relatively low (Hoeksema, 

Naturalis Biodiversity Center, personal 

communication, 2013) and the ReefBase 

coral bleaching database only reports 

events with low bleaching severity as 

the worst past events to ever occur 

there. We have found no speciesspecific information on the 

susceptibility of this species to 

bleaching or ocean acidification; 

however, its growth form suggests it is 

not among the most susceptible species 

(Wooldridge, 2014). 

Anthropogenic eutrophication occurs 

in the range of the species near the 

capital of Noumea and is attributed 

mostly to inadequately treated sewage 

(Fichez et al., 2010), although 19 

aquaculture farms on the west coast and 

island-wide agriculture may also play 

roles (David et al., 2010). Storm events 

and flooding have also recently 

occurred in the range of the species 

(EMR, 2013), and there is concern that 

climate change may make such events 

more frequent in New Caledonia 

(Gilbert, personal communication, 

2013). 

The biggest threats to New Guinea’s 

coral reef resources include 

sedimentation and pollution from 

inland sources (e.g., forest clearance, 

sewage, and erosion), climate change, 

and dynamite fishing (Burke et al., 2011; 

PNG, 2009; PNG, 2012). There is little 

specific data on these threats in New 

Guinea in the above references. 

Siderastrea glynni 

Should S. glynni ever be restored to 

the wild, it faces considerable habitat 

degradation threats from coastal 

development, oil production, 

sedimentation, eutrophication and other 

pollution, and increased transportation 

activities in the Panama City area, the 

Gulf of Panama, and the enlarged 

Panama Canal, which is due to open in 

2016 (Mate, 2003; Guzma ́n and Edgar, 

2008). Almost continuous dredging and 

release of oil-based compounds (bunker 

oil, diesel, gasoline, etc.) that are spilled 

from nearby port facilities and 

commercial vessels anchored near the 

species’ natural range are other reasons 

why it was decided to transfer and then 

keep in captivity the remaining known 

colonies (Guzman, personal 

communication, 2013). ‘‘During the 

1997–98 ENSO event, the four known 

colonies of S. glynni began to 

deteriorate, displaying bleaching and 

tissue loss’’ (Guzma ́n and Edgar, 2008). 

This suggests this species is vulnerable 

to increased ocean temperatures, though 

there is no specific research on this 

point. As discussed above, the area of 

the species’ range is subject to a high 

frequency of bleaching warnings. We 

have found no species-specific 

information on the susceptibility of this 

species to ocean acidification. 

Tubastraea floreana 

For T. floreana, there is a lack of 

information on thermal tolerances, but 

‘‘the dramatic reduction in its 

distribution immediately after the 1982– 

83 [ENSO] event suggests that this 

mortality resulted from the event’’ 

(Hickman et al., 2007). This is true 

despite the fact that this species is 

azooxanthellate, suggesting that other 

mechanisms besides loss of calorie 

subsidy from symbionts are involved. 

Edgar et al. (2010) document a series of 

drastic ecosystem changes in the 

Galapagos following the 1982–83 ENSO 

event, including dramatic declines in 

dissolved nutrients and phytoplankton 

productivity, leading to declines across 

the food chain and resulting in heavily 

grazed reefs with crustose coralline 

algae (‘‘urchin barrens’’) replacing 

former macroalgal and coral habitats. A 

total of 95–99 percent of reef coral cover 

was lost from the Galapagos between 

1983 and 1985 (Edgar et al., 2010). All 

known coral reefs based on calcareous 

frameworks died and subsequently 

disintegrated to rubble and sand (Glynn, 

1994). These changes led to large 

decreases in biodiversity. The urchin 

Eucidaris galapagensis now appears to 

be present in sufficient numbers to 

prevent re-establishment of coral and 

macroalgal habitat, thereby facilitating a 

regime shift in local benthic habitats 

(Edgar et al., 2010). Moreover, the 

Galapagos Islands sit near the center of 

the most intense El Nin ̃ o events in the 

region (Glynn and Ault, 2000) and are 

regularly included in bleaching threat 

warnings issued by NOAA (see above). 

Therefore, future ENSO events and 

inhibition of recruitment are likely to 

remain threats to T. floreana. We have 

found no species-specific information 

on the susceptibility of this species to 

ocean acidification. 

Disease and Predation 

Coral disease has been linked to the 

effects of climate change (see Brainard 

et al., 2011), especially indirectly as a 

synergistic effect, as climate change and 

other threats potentially increase stress 

on corals, making them more 

susceptible to disease. Coral diseases 

also appear to be increasing worldwide 

(Roessig et al., 2004). Nevertheless, 

susceptibility of coral species to disease 

is highly species-specific and no 

generalizations can be made. We found 

no species-specific information on 

disease in C. noumeae or T. floreana. 

Black-band, dark spot, and white plague 

diseases in the Caribbean occur in S. 

siderea, which is closely related to S. 

glynni (Sekar et al., 2008; Brandt and 

McManus, 2009; Cardenas et al., 2012), 

suggesting S. glynni may be susceptible 

to similar coral diseases, but we have no 

solid information. 

With respect to predation, we found 

no information on predation threats to 

S. glynni or T. floreana. For C. noumeae, 

one potential predation threat is 

Acanthaster planci (crown-of-thorns 

starfish). Acanthaster planci does not 

appear to be a major cause of coral 

mortality in New Caledonia (Adjeroud, 

2012), but several remote reefs surveyed 

during the Global Reef Expedition in 

November 2013 on the outer-slope of 

Guilbert’s atolls showed evidence of 

past outbreaks (LOF, 2013). 

Inadequacy of Existing Regulatory 

Mechanisms 

The petitioners discussed regulation 

of trade in corals under CITES as a 

threat to these species. All of the species 

considered in this petition were listed 

in Appendix II of CITES in 1989, when 

all scleractinian corals were listed. 

While only some scleractinians were in 

trade at the time, the 1989 listing 

rationale for including all scleractinians 

in Appendix II was because of 

identification difficulties where nontraded species resemble species in trade. 

According to Article II of CITES, species 

listed on Appendix II are those that are 

‘‘not necessarily now threatened with 

extinction but may become so unless 

trade in specimens of such species is 

subject to strict regulation in order to 

avoid utilization incompatible with 

their survival.’’ Based on the CITES 

definitions and standards for listing 

species on Appendix II, the species’ 

listing on Appendix II is not itself an 

inherent indication that these species 

may now warrant threatened or 

endangered status under the ESA. The 

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significance of any threat from 

international trade would depend on the 

amount of international trade relative to 

the population size of the species, as 

well as any other factors related to the 

trade, such as habitat damage caused in 

the collecting process, or synergistic 

effects of other threats. We have no 

information any of these three species is 

traded internationally. 

Because each of the species 

considered herein exists in small ranges 

that do not overlap with each other, and 

they are not otherwise managed or 

regulated under any other common 

international regimes, additional 

discussion of this factor is left for the 

species-specific entries for this section, 

below. 

Cantharellus noumeae 

Since the Organic Law (No. 99–209) 

on March 19, 1999, New Caledonia has 

been recognized as an ‘‘Overseas 

Country’’ of France. This status gives 

New Caledonia extensive autonomy 

with respect to France. In particular, the 

national laws in force within France are 

no longer applicable to New Caledonia, 

and New Caledonia now manages the 

ocean resources of its Exclusive 

Economic Zone. The territorial sea and 

the maritime public domain (coastal 

terrestrial and nearshore aquatic zone 

originating under French colonial law) 

depend on management from New 

Caledonia’s three provinces (David et 

al., 2010). In the two provinces where C. 

noumeae occurs, collection of live 

corals (and other marine resources) is 

restricted to scientists and licensed 

fishers who can only collect for a 

domestic market. 

The range of C. noumeae is included 

in the United Nations Education, 

Scientific and Cultural Organization 

(UNESCO) World Heritage Site 

designation for the ‘‘Lagoons of New 

Caledonia’’ site, specifically within the 

South Grand Lagoon area. The World 

Heritage Site implementation is 

supported by specific legislation on 

fisheries, land and water use planning, 

urban development, and mining (Morris 

and Mackay, 2008). A wide monitoring 

program of the heritage site all around 

New Caledonia was created (Andre ́foue ̈t 

2008), but this suffers from a lack of 

sampling at a species level (Gilbert, 

personal communication, 2013). In 

2011, the World Heritage Committee of 

UNESCO (the organizing body for World 

Heritage Sites) issued Decision 35Com 

7B.22, which expressed concern 

regarding permits granted to the mining 

company GEOVIC to explore for cobalt 

in mineral sands in areas adjacent to the 

site and near the range of C. noumeae. 

The committee requested that New 

Caledonia submit Environmental Impact 

Assessments for the proposed 

exploration and possible exploitation of 

cobalt sands to the World Heritage 

Centre. We have no evidence this has 

occurred. The New Caledonian Mining 

Code prescribes mitigation measures to 

mitigate the impacts of mining activities 

(see above), and abandoned mines are 

being restored using indigenous plant 

species (UNESCO, 2011). 

In Papua New Guinea, there is a 

variety of legislation to protect 

biodiversity and habitat, including a 

mandate to ensure marine resource 

sustainability, and a plan of action 

directed at coral reef conservation (PNG, 

2009). However, as noted above, threats 

remain. Resources and capacity may not 

be adequate to ensure full 

implementation of the laws and plan 

(PNG, 2009; PNG, 2012). 

Overall, we do not believe that the 

threat to C. noumeae from habitat 

modification, destruction, and pollution 

is adequately addressed or mitigated by 

existing regulatory mechanisms. 

Siderastrea glynni 

A national law in Panama prohibits 

coral extraction or mining (Guzma ́n, 

2003), but enforcement is weak and the 

law may not fully protect rare species 

(Guzma ́n, personal communication, 

2013). The range of S. glynni is adjacent 

to the Bay of Panama, which is 

designated an internationally important 

wetland under the Ramsar Convention 

and contains extensive mangrove beds 

that are critical nursery grounds for 

many marine species. The Bay is a 

protected Wildlife Refuge under 

Panamanian law. However, developers 

seek to open the area for tourism, and 

Panamanian authorities have requested 

a reduction of the Ramsar area of the 

bay (AIDA, 2013). We were not able to 

find any other species-specific 

information on this threat. Based on the 

available information, it is not clear that 

existing regulatory mechanisms would 

be adequate to protect S. glynni, should 

it be reintroduced into the wild or found 

in additional locations. 

Tubastraea floreana 

The Gala ́pagos Marine Reserve was 

established in 1986 and expanded to its 

current size around all the islands in 

1998. The reserve has a zoning plan 

with both limited and multiple use 

zones. Rules prohibit removing or 

disturbing any plant, animal, or remains 

of such, or other natural objects. 

Tubastraea floreana also occurs inside 

the Galapagos Island World Heritage 

Site (expanded to include Galapagos 

Marine Reserve areas in 2001) and the 

Gala ́pagos Island Man and Biosphere 

Reserve (1984), both designations of 

UNESCO. The area was also designated 

a Gala ́pagos Archipelago Particularly 

Sensitive Area in 2005. This is a 

designation by the International 

Maritime Organization (IMO) that 

recognizes the area as having ecological, 

socio-economic, or scientific attributes 

that make the area vulnerable to damage 

by international shipping activities. 

Based on this designation, the IMO 

instituted special navigation rules in the 

area. In addition, Ecuador’s ‘‘Ley de 

Gestion Ambiental’’ (Law of 

Environmental Management) establishes 

principles and directives for 

environmental management, land-use 

planning, zoning, sustainable use, and 

natural heritage conservation. Ecuador’s 

fisheries law states that no harm may be 

caused to areas that are declared 

protected, with corals included under 

those protections (MCA Toolkit, 2013). 

While the above laws and protected area 

designations provide a great deal of 

protection for resources in the area in 

principal, in practice, illegal activities 

and incomplete and difficult 

enforcement, as discussed in the status 

review report (Meadows, 2014), could 

threaten T. floreana. Moreover, the 

threats from climate change and ENSO 

events are outside the scope of these 

protections. 

Other Natural or Manmade Factors 

Affecting Their Continued Existence 

The range of C. noumeae in New 

Caledonia is exposed to eight tropical 

storms per year on average (David et al., 

2010). Specific effects of storms on this 

species are not documented, but the 

petitioner submitted an undated Web 

page that claims Cyclone Erica 

destroyed between 10 and 80 percent of 

live coral in New Caledonia in 2003 

(EDGE, Undated; Guillemot et al., 2010). 

We were not able to find any other 

species-specific information available 

regarding this threat category for C. 

noumeae. Based on this information, we 

consider tropical storms an additional 

potential natural threat to the species, 

for which we seek additional 

information (see below). 

For S. glynni and T. floreana, both 

species have such a small number of 

colonies, they are susceptible to all of 

the problems of species with low 

genetic diversity and population size, 

including inbreeding depression, 

population bottlenecks, Allee effects, 

and density-independent mortality, 

among others. 

Extinction Risk 

The extinction risk analyses of 

Meadows (2014) found all three species 

to be at either a moderately high or high 

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risk of extinction. The extinction risk for 

C. noumeae was found to be moderately 

high, based on the species’ small, 

restricted range, likely low growth rate 

and genetic diversity, and potential 

threats from development, water 

pollution, possibly sedimentation at 

some level, and potential illegal 

activities, mitigated by consideration of 

potential resilience to sedimentation 

threats and uncertainty regarding 

sensitivity to heavy metals. Based on the 

current information, this is the case 

whether or not the species’ range 

includes New Guinea. The extinction 

risk for S. glynni was found to be high, 

due to the lack of known populations in 

the wild, a small captive population in 

a single location, likely low growth rates 

and genetic diversity, and potential 

increased threats from El Nin ̃ o, climate 

change, disease, and other development 

and habitat degradation, should the 

species be reintroduced to Panama. The 

extinction risk for T. floreana was found 

to be high, based on its small, restricted 

range, documented declines, likely low 

levels of genetic diversity, and threats 

from El Nin ̃ o, climate change, 

development, and illegal activities, 

mitigated by potential for moderate 

productivity. 

After reviewing the best available 

scientific data and the extinction risk 

evaluations of the three species of coral, 

we concur with Meadows (2014) and 

conclude that the risk of extinction for 

all three species is currently high. 

Protective Efforts 

We evaluated conservation efforts we 

are aware of to protect and recover coral 

that are either underway but not yet 

shown to be effective, or are only 

planned. We were not able to find any 

information on conservation efforts 

specific to C. noumeae or T. floreana, or 

their habitat, that are not yet 

implemented or shown to be effective 

and that would potentially alter the 

extinction risk for the species. For S. 

glynni, we are aware that Dr. Hector 

Guzma ́n, who maintains the only 

surviving colonies of this species in 

captivity at the STRI laboratories, is 

planning to cryopreserve some 

specimens to provide an additional 

means to recover the species and 

preserve its genetic information. The 

certainty that this effort will be 

implemented is unclear. Further, the 

effectiveness of a cryopreservation effort 

for species recovery is largely unknown, 

and thus it is impossible to determine 

whether this effort will be effective in 

conserving or improving the status of 

this species. We are thus not able to 

conclude that any current conservation 

efforts would alter the extinction risk for 

any of these three species. We seek 

additional information on other 

conservation efforts in our public 

comment process (see below). 

Proposed Determination 

Based on the best available scientific 

and commercial information as 

presented in the status report and this 

finding, we find that all three species of 

coral are in danger of extinction 

throughout all of their ranges. We 

assessed the ESA section 4(a)(1) factors 

and conclude that Cantharellus 

noumeae, Siderastrea glynni, and 

Tubastraea floreana all face ongoing 

threats from habitat alteration, small 

ranges and/or population sizes, and the 

inadequacy of existing regulatory 

mechanisms throughout their ranges. C. 

noumeae also faces risks from pollution 

and S. glynni may be at risk from 

disease. We therefore propose to list all 

three species as endangered. 

Effects of Listing 

Conservation measures provided for 

species listed as endangered or 

threatened under the ESA include 

recovery actions (16 U.S.C. 1533(f)); 

concurrent designation of critical 

habitat, if prudent and determinable (16 

U.S.C. 1533(a)(3)(A)); Federal agency 

requirements to consult with NMFS 

under section 7 of the ESA to ensure 

their actions do not jeopardize the 

species or result in adverse modification 

or destruction of critical habitat should 

it be designated (16 U.S.C. 1536); and 

prohibitions on taking (16 U.S.C. 1538). 

Recognition of the species’ plight 

through listing promotes conservation 

actions by Federal and state agencies, 

foreign entities, private groups, and 

individuals. The main effects of the 

proposed endangered listings are 

prohibitions on take, including export 

and import. 

Identifying Section 7 Conference and 

Consultation Requirements 

Section 7(a)(2) (16 U.S.C. 1536(a)(2)) 

of the ESA and NMFS/USFWS 

regulations require Federal agencies to 

consult with us to ensure that activities 

they authorize, fund, or carry out are not 

likely to jeopardize the continued 

existence of listed species or destroy or 

adversely modify critical habitat. 

Section 7(a)(4) (16 U.S.C. 1536(a)(4)) of 

the ESA and NMFS/USFWS regulations 

also require Federal agencies to confer 

with us on actions likely to jeopardize 

the continued existence of species 

proposed for listing, or that result in the 

destruction or adverse modification of 

proposed critical habitat of those 

species. It is unlikely that the listing of 

these species under the ESA will 

increase the number of section 7 

consultations, because these species 

occur outside of the United States and 

are unlikely to be affected by Federal 

actions. 

Critical Habitat 

Critical habitat is defined in section 3 

of the ESA (16 U.S.C. 1532(5)) as: (1) 

The specific areas within the 

geographical area occupied by a species, 

at the time it is listed in accordance 

with the ESA, on which are found those 

physical or biological features (a) 

essential to the conservation of the 

species and (b) that may require special 

management considerations or 

protection; and (2) specific areas outside 

the geographical area occupied by a 

species at the time it is listed upon a 

determination that such areas are 

essential for the conservation of the 

species. ‘‘Conservation’’ means the use 

of all methods and procedures needed 

to bring the species to the point at 

which listing under the ESA is no 

longer necessary. Section 4(a)(3)(A) of 

the ESA (16 U.S.C. 1533(a)(3)(A)) 

requires that, to the extent prudent and 

determinable, critical habitat be 

designated concurrently with the listing 

of a species. However, critical habitat 

shall not be designated in foreign 

countries or other areas outside U.S. 

jurisdiction (50 CFR 424.12 (h)). 

The best available scientific and 

commercial data as discussed above 

identify the geographical areas occupied 

by Aipysurus fuscus, Cantharellus 

noumeae, Centrophorus harrissoni, 

Pterapogon kauderni, Siderastrea 

glynni, and Tubastraea floreana as being 

entirely outside U.S. jurisdiction, so we 

cannot designate critical habitat for 

these species. 

We can designate critical habitat in 

areas in the United States currently 

unoccupied by the species, if the area(s) 

are determined by the Secretary to be 

essential for the conservation of the 

species. Regulations at 50 CFR 424.12(e) 

specify that we shall designate as 

critical habitat areas outside the 

geographical range presently occupied 

by the species only when the 

designation limited to its present range 

would be inadequate to ensure the 

conservation of the species. The best 

available scientific and commercial 

information on these species does not 

indicate that U.S. waters provide any 

specific essential biological function for 

any of the species proposed for listing. 

Based on the best available information, 

we have not identified unoccupied 

area(s) in U.S. water that are currently 

essential to the conservation of any of 

the corals proposed for listing. 

Therefore, based on the available 

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74982 Federal Register / Vol. 79, No. 241 / Tuesday, December 16, 2014 / Proposed Rules 

information, we do not intend to 

designate critical habitat for Aipysurus 

fuscus, Cantharellus noumeae, 

Centrophorus harrissoni, Pterapogon 

kauderni, Siderastrea glynni, and 

Tubastraea floreana. 

Identification of Those Activities That 

Would Constitute a Violation of Section 

9 of the ESA 

On July 1, 1994, NMFS and FWS 

published a policy (59 FR 34272) that 

requires us to identify, to the maximum 

extent practicable at the time a species 

is listed, those activities that would or 

would not constitute a violation of 

section 9 of the ESA. 

Because we are proposing to list all 

three corals and the dusky sea snake as 

endangered, all of the prohibitions of 

section 9(a)(1) of the ESA will apply to 

these species. These include 

prohibitions against the import, export, 

use in foreign commerce, or ‘‘take’’ of 

the species. These prohibitions apply to 

all persons subject to the jurisdiction of 

the United States, including in the 

United States, its territorial sea, or on 

the high seas. Take is defined as ‘‘to 

harass, harm, pursue, hunt, shoot, 

wound, kill, trap, capture, or collect, or 

to attempt to engage in any such 

conduct.’’ The intent of this policy is to 

increase public awareness of the effects 

of this listing on proposed and ongoing 

activities within the species’ range. 

Activities that we believe could result in 

a violation of section 9 prohibitions for 

these species include, but are not 

limited to, the following: 

(1) Possessing, delivering, 

transporting, or shipping any individual 

or part (dead or alive) taken in violation 

of section 9(a)(1); 

(2) Delivering, receiving, carrying, 

transporting, or shipping in interstate or 

foreign commerce any individual or 

part, in the course of a commercial 

activity; 

(3) Selling or offering for sale in 

interstate commerce any part, except 

antique articles at least 100 years old; 

(4) Importing or exporting; 

(5) Releasing captive animals into the 

wild without a permit issued under 

section 10(a)(1)(A). Although animals 

held non-commercially in captivity at 

the time of listing are exempt from the 

prohibitions of import and export, the 

individual animals are considered listed 

and afforded most of the protections of 

the ESA, including most importantly, 

the prohibition against injuring or 

killing. Release of a captive animal has 

the potential to injure or kill the animal. 

Of an even greater conservation 

concern, the release of a captive animal 

has the potential to affect wild 

populations through introduction of 

diseases or inappropriate genetic 

mixing; 

(6) Harming captive animals by, 

among other things, injuring or killing a 

captive animal, through experimental or 

potentially injurious care or conducting 

research or sexual breeding activities on 

captive animals, outside the bounds of 

normal animal husbandry practices. 

Captive sexual breeding of corals is 

considered potentially injurious. 

Furthermore, the production of coral 

progeny has conservation implications 

(both positive and negative) for wild 

populations. Experimental or 

potentially injurious care or procedures 

and research or sexual breeding 

activities of corals or dusky sea snakes 

may, depending on the circumstances, 

be authorized under an ESA 10(a)(1)(A) 

permit for scientific research or the 

enhancement of the propagation or 

survival of the species. 

Identification of Those Activities That 

Would Not Constitute a Violation of 

Section 9 of the ESA 

We will identify, to the extent known 

at the time of the final rule, specific 

activities that will not be considered 

likely to result in a violation of section 

9 of the ESA. Although not binding, we 

are considering the following actions, 

depending on the circumstances, as not 

being prohibited by ESA section 9: 

(1) Take authorized by, and carried 

out in accordance with the terms and 

conditions of, an ESA section 

10(a)(1)(A) permit issued by NMFS for 

purposes of scientific research or the 

enhancement of the propagation or 

survival of the species; 

(2) Continued possession of parts that 

were in possession at the time of listing. 

Such parts may be non-commercially 

exported or imported; however the 

importer or exporter must be able to 

provide evidence to show that the parts 

meet the criteria of ESA section 9(b)(1) 

(i.e., held in a controlled environment at 

the time of listing, in a non-commercial 

activity); 

(3) Continued possession of live 

corals or sea snakes that were in 

captivity or in a controlled environment 

(e.g., in aquaria) at the time of this 

listing, so long as the prohibitions under 

ESA section 9(a)(1) are not violated. 

Facilities must provide evidence that 

the animals were in captivity or in a 

controlled environment prior to listing. 

We suggest such facilities submit 

information to us on the animals in their 

possession (e.g., size, age, description of 

animals, and the source and date of 

acquisition) to establish their claim of 

possession (see FOR FURTHER

INFORMATION CONTACT); 

(4) Provision of care for live corals or 

sea snakes that were in captivity at the 

time of listing. These individuals are 

still protected under the ESA and may 

not be killed or injured, or otherwise 

harmed, and, therefore, must receive 

proper care. Normal care of captive 

animals necessarily entails handling or 

other manipulation of the animals, and 

we do not consider such activities to 

constitute take or harassment of the 

animals so long as adequate care, 

including veterinary care, when such 

practices, procedures, or provisions are 

not likely to result in injury, is 

provided; and 

(5) Any interstate and foreign 

commerce trade of animals already in 

captivity. Section 11(f) of the ESA gives 

NMFS authority to promulgate 

regulations that may be appropriate to 

enforce the ESA. NMFS may promulgate 

future regulations to regulate trade or 

holding of these species (if any), if 

necessary. NMFS will provide the 

public with the opportunity to comment 

on future proposed regulations. 

Protective Regulations Under Section 

4(d) of the ESA 

We are proposing to list Pterapogon 

kauderni, and Centrophorus harrissoni 

as threatened species. In the case of 

threatened species, ESA section 4(d) 

leaves it to the Secretary’s discretion 

whether, and to what extent, to extend 

the section 9(a) ‘‘take’’ prohibitions to 

the species, and authorizes us to issue 

regulations necessary and advisable for 

the conservation of the species. Thus, 

we have flexibility under section 4(d) to 

tailor protective regulations, taking into 

account the effectiveness of available 

conservation measures. The 4(d) 

protective regulations may prohibit, 

with respect to threatened species, some 

or all of the acts which section 9(a) of 

the ESA prohibits with respect to 

endangered species. These 9(a) 

prohibitions apply to all individuals, 

organizations, and agencies subject to 

U.S. jurisdiction. We will consider 

potential protective regulations 

pursuant to section 4(d) for the 

proposed threatened species. For 

example, we may consider future 

regulations on trade of wild-caught 

Banggai cardinalfish (see number 7 

below). We seek public comment on 

potential 4(d) protective regulations (see 

below). 

Public Comments Solicited 

To ensure that any final action 

resulting from this proposed rule to list 

six species will be as accurate and 

effective as possible, we are soliciting 

comments and information from the 

public, other concerned governmental 

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agencies, the scientific community, 

industry, and any other interested 

parties on information in the status 

review and proposed rule. Comments 

are encouraged on these proposals (See 

DATES and ADDRESSES). We must base 

our final determination on the best 

available scientific and commercial 

information when making listing 

determinations. We cannot, for example, 

consider the economic effects of a 

listing determination. Final 

promulgation of any regulation(s) on 

these species’ listing proposals will take 

into consideration the comments and 

any additional information we receive, 

and such communications may lead to 

a final regulation that differs from this 

proposal or result in a withdrawal of 

this listing proposal. We particularly 

seek: 

(1) Information concerning the threats 

to any of the six species proposed for 

listing; 

(2) Taxonomic information on any of 

these species; 

(3) Biological information (life 

history, genetics, population 

connectivity, etc.) on any of these 

species; 

(4) Efforts being made to protect any 

of these species throughout their current 

ranges; 

(5) Information on the commercial 

trade of any of these species; 

(6) Historical and current distribution 

and abundance and trends for any of 

these species; and 

(7) Information relevant to potential 

ESA section 4(d) protective regulations 

for any of the proposed threatened 

species, especially the application, if 

any, of the ESA section 9 prohibitions 

on import, take, possession, receipt, and 

sale of the Banggai cardinalfish which is 

currently in international trade. 

We request that all information be 

accompanied by: (1) Supporting 

documentation, such as maps, 

bibliographic references, or reprints of 

pertinent publications; and (2) the 

submitter’s name, address, and any 

association, institution, or business that 

the person represents. 

Role of Peer Review 

In December 2004, the Office of 

Management and Budget (OMB) issued 

a Final Information Quality Bulletin for 

Peer Review establishing a minimum 

peer review standard. Similarly, a joint 

NMFS/FWS policy (59 FR 34270; July 1, 

1994) requires us to solicit independent 

expert review from qualified specialists, 

concurrent with the public comment 

period. The intent of the peer review 

policy is to ensure that listings are based 

on the best scientific and commercial 

data available. We solicited peer review 

comments on each of the status review 

reports, including from: four scientists 

with expertise on sea snakes or the 

dusky sea snake specifically, five 

familiar with the Banggai cardinalfish, 

five familiar with Harrisson’s dogfish, 

and ten scientists familiar with corals. 

For these species, we received 

comments from the scientists, and their 

comments are incorporated into the 

draft status review reports for each 

species and this 12-month finding. 

Proposed Revisions to the NMFS Lists 

We propose to revise and add table 

subheadings to accommodate the 

proposed listings in our lists of 

threatened and endangered species at 50 

CFR 223.102 and 50 CFR 224.101, 

respectively. We propose to revise the 

subheading of ‘‘Sea Turtles’’ in both 

tables by changing the subheading to 

‘‘Reptiles.’’ This new subheading will 

encompass all currently listed sea 

turtles as well as other marine reptiles 

like the dusky sea snake. In addition, we 

propose to add the subheading ‘‘Corals’’ 

to our table at 50 CFR 224.101. This 

subheading has already been added to 

our table at 50 CFR 223.102 in a 

previous rulemaking (79 FR 20802; 

April 14, 2014). These revisions and 

additions are not substantive changes, 

but having these headings will help the 

public identify and locate species of 

interest in a more efficient manner. 

References 

A complete list of the references used 

in this proposed rule is available upon 

request (see ADDRESSES). 

Classification 

National Environmental Policy Act 

The 1982 amendments to the ESA, in 

section 4(b)(1)(A), restrict the 

information that may be considered 

when assessing species for listing. Based 

on this limitation of criteria for a listing 

decision and the opinion in Pacific 

Legal Foundation v. Andrus, 675 F. 2d 

825 (6th Cir. 1981), NMFS has 

concluded that ESA listing actions are 

not subject to the environmental 

assessment requirements of the National 

Environmental Policy Act (NEPA) (See 

NOAA Administrative Order 216–6). 

Executive Order 12866, Regulatory 

Flexibility Act, and Paperwork 

Reduction Act 

As noted in the Conference Report on 

the 1982 amendments to the ESA, 

economic impacts cannot be considered 

when assessing the status of a species. 

Therefore, the economic analysis 

requirements of the Regulatory 

Flexibility Act are not applicable to the 

listing process. In addition, this 

proposed rule is exempt from review 

under Executive Order 12866. This 

proposed rule does not contain a 

collection-of-information requirement 

for the purposes of the Paperwork 

Reduction Act. 

Executive Order 13132, Federalism 

In accordance with E.O. 13132, we 

determined that this proposed rule does 

not have significant Federalism effects 

and that a Federalism assessment is not 

required. In keeping with the intent of 

the Administration and Congress to 

provide continuing and meaningful 

dialogue on issues of mutual state and 

Federal interest, this proposed rule will 

be given to the relevant governmental 

agencies in the countries in which the 

species occurs, and they will be invited 

to comment. We will confer with the 

U.S. Department of State to ensure 

appropriate notice is given to foreign 

nations within the range of all three 

species. As the process continues, we 

intend to continue engaging in informal 

and formal contacts with the U.S. State 

Department, giving careful 

consideration to all written and oral 

comments received. 

List of Subjects in 50 CFR Parts 223 and 

224 

Administrative practice and 

procedure, Endangered and threatened 

species, Exports, Imports, Reporting and 

record keeping requirements, 

Transportation. 

Dated: December 8, 2014. 

Samuel D. Rauch, III. 

Deputy Assistant Administrator for 

Regulatory Programs, National Marine 

Fisheries Service. 

For the reasons set out in the 

preamble, 50 CFR parts 223 and 224 are 

proposed to be amended as follows: 

PART 223—THREATENED MARINE 

AND ANADROMOUS SPECIES 

■ 1. The authority citation for part 223 

continues to read as follows: 

Authority: 16 U.S.C. 1531–1543; subpart B, 

§ 223.201–202 also issued under 16 U.S.C. 

1361 et seq.; 16 U.S.C. 5503(d) for 

§ 223.206(d)(9). 

■ 2. In § 223.102, amend the table in 

paragraph (e) by: 

■ A. Revising the table subheading of 

‘‘Sea Turtles 2’’ to ‘‘Reptiles 2’’; and 

■ B. Adding new entries for two species 

in alphabetical order under the ‘‘Fishes’’ 

table subheading to read as follows: 

§ 223.102 Enumeration of threatened 

marine and anadromous species. 

* * * * * 

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74984 Federal Register / Vol. 79, No. 241 / Tuesday, December 16, 2014 / Proposed Rules 

(e) The threatened species under the 

jurisdiction of the Secretary of 

Commerce are: 

Species 1

Citation(s) for listing 

determination(s) 

Critical 

habitat ESA rules 

Common name Scientific name Description of listed entity 

*******

REPTILES 2

* * * * * * * 

FISHES

Cardinalfish, Banggai ......... Pterapogon kauderni ........ Entire species ................... Insert Federal Register citation and date when 

published as a final rule].

NA NA 

* * * * * * * 

Shark, Harrisson’s dogfish Centrophorus harrissoni ... Entire species ................... Insert Federal Register citation and date when 

published as a final rule].

NA NA 

1Species includes taxonomic species, subspecies, distinct population segments (DPSs) (for a policy statement, see 61 FR 4722, February 7, 

1996), and evolutionarily significant units (ESUs) (for a policy statement, see 56 FR 58612, November 20, 1991). 2 Jurisdiction for sea turtles by the Department of Commerce, National Oceanic and Atmospheric Administration, National Marine Fisheries 

Service, is limited to turtles while in the water. 

PART 224—ENDANGERED MARINE 

AND ANADROMOUS SPECIES 

■ 3. The authority citation for part 224 

continues to read as follows: 

Authority: 16 U.S.C. 1531–1543 and 16 

U.S.C. 1361 et seq. 

■ 4. In § 224.101, paragraph (h), amend 

the table by: 

■ A. Revising the table subheading of 

‘‘Sea Turtles 2’’ to ‘‘Reptiles 2’’; 

■ B. Adding an entry for the dusky sea 

snake in alphabetical order under the 

new ‘‘Reptiles 2’’ table subheading; 

■ C. Adding a ‘‘Corals’’ table 

subheading to follow the ‘‘Molluscs’’ 

table subheading; and 

■ D. Adding entries for three species of 

coral in alphabetical order by scientific 

name under the ‘‘Corals’’ table 

subheading to read as follows: 

§ 224.101 Enumeration of endangered 

marine and anadromous species. 

* * * * * 

(h) The endangered species under the 

jurisdiction of the Secretary of 

Commerce are: 

Species 1

Citation(s) for listing 

determination(s) 

Critical 

habitat ESA rules 

Common name Scientific name Description of listed entity 

*******

REPTILES 2

Sea snake, dusky .............. Aipysurus fuscus ............... Entire species ................... Insert Federal Register citation and date when 

published as a final rule].

NA NA 

* * * * * * * 

MOLLUSCS

* * * * * * * 

CORALS

Coral, [no common name] Cantharellus noumeae ...... Entire species ................... Insert Federal Register citation and date when 

published as a final rule].

NA NA 

Coral, [no common name] Siderastrea glynni ............. Entire species ................... Insert Federal Register citation and date when 

published as a final rule].

NA NA 

Coral, [no common name] Tubastraea floreana .......... Entire species ................... Insert Federal Register citation and date when 

published as a final rule].

NA NA 

1Species includes taxonomic species, subspecies, distinct population segments (DPSs) (for a policy statement, see 61 FR 4722, February 7, 

1996), and evolutionarily significant units (ESUs) (for a policy statement, see 56 FR 58612, November 20, 1991). 2 Jurisdiction for sea turtles by the Department of Commerce, National Oceanic and Atmospheric Administration, National Marine Fisheries 

Service, is limited to turtles while in the water. 

* * * * * 

[FR Doc. 2014–29203 Filed 12–15–14; 8:45 am] 

BILLING CODE 3510–22–P 

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