Proposed Rule2022-19109
Endangered and Threatened Wildlife and Plants: Proposed Rule to List the Queen Conch as Threatened Under the Endangered Species Act (ESA)
Primary source
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Published
September 8, 2022
Issuing agencies
Commerce DepartmentNational Oceanic and Atmospheric Administration
Abstract
We, NMFS, announce a proposed rule to list the queen conch (Aliger gigas, previously known as Strombus gigas) as a threatened species under the Endangered Species Act (ESA). We have completed a comprehensive status review for the queen conch. After considering the status review report, and after taking into account efforts being made to protect the species, we have determined that the queen conch is likely to become an endangered species within the foreseeable future throughout its range. Therefore, we propose to list the queen conch as a threatened species under the ESA. Any protective regulations determined to be necessary and advisable for the conservation of the queen conch under ESA would be proposed in a subsequent Federal Register announcement. We solicit information to assist this listing determination, the development of proposed protective regulations, and designation of critical habitat within U.S jurisdiction.
Full Text
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[Federal Register Volume 87, Number 173 (Thursday, September 8, 2022)]
[Proposed Rules]
[Pages 55200-55239]
From the Federal Register Online via the Government Publishing Office [<a href="http://www.gpo.gov">www.gpo.gov</a>]
[FR Doc No: 2022-19109]
[[Page 55199]]
Vol. 87
Thursday,
No. 173
September 8, 2022
Part IV
Department of Commerce
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National Oceanic and Atmospheric Administration
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50 CFR Part 223
Endangered and Threatened Wildlife and Plants: Proposed Rule To List
the Queen Conch as Threatened Under the Endangered Species Act (ESA);
Proposed Rule
Federal Register / Vol. 87 , No. 173 / Thursday, September 8, 2022 /
Proposed Rules
[[Page 55200]]
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DEPARTMENT OF COMMERCE
National Oceanic and Atmospheric Administration
50 CFR Part 223
[Docket No. 220830-0177; RTID 0648-XR071]
Endangered and Threatened Wildlife and Plants: Proposed Rule to
List the Queen Conch as Threatened Under the Endangered Species Act
(ESA)
AGENCY: National Marine Fisheries Service (NMFS), National Oceanic and
Atmospheric Administration (NOAA), Commerce.
ACTION: Proposed rule; request for comments.
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SUMMARY: We, NMFS, announce a proposed rule to list the queen conch
(Aliger gigas, previously known as Strombus gigas) as a threatened
species under the Endangered Species Act (ESA). We have completed a
comprehensive status review for the queen conch. After considering the
status review report, and after taking into account efforts being made
to protect the species, we have determined that the queen conch is
likely to become an endangered species within the foreseeable future
throughout its range. Therefore, we propose to list the queen conch as
a threatened species under the ESA. Any protective regulations
determined to be necessary and advisable for the conservation of the
queen conch under ESA would be proposed in a subsequent Federal
Register announcement. We solicit information to assist this listing
determination, the development of proposed protective regulations, and
designation of critical habitat within U.S jurisdiction.
DATES: Information and comments on this proposed rule must be received
by November 7, 2022. Public hearing requests must be requested by
October 24, 2022.
ADDRESSES: You may submit comments, information, or data on this
document, identified by the code NOAA-NMFS-2019-0141 by any of the
following methods:
<bullet> Electronic Submissions: Submit all electronic comments via
the Federal eRulemaking Portal. Go to <a href="http://www.regulations.gov">www.regulations.gov</a> and enter
NOAA-NMFS-2019-0141 in the Search box. Click on the ``Comment'' icon,
complete the required fields, and enter or attach your comments.
<bullet> Mail: NMFS, Southeast Regional Office, 263 13th Avenue
South, St. Petersburg, FL 33701;
<bullet> Instructions: Comments sent by any other method, to any
other address or individual, or received after the end of the comment
period, might not be considered by NMFS. All comments received are a
part of the public record and will generally be posted for public
viewing on <a href="http://www.regulations.gov">www.regulations.gov</a> 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. NMFS will accept anonymous
comments (enter ``N/A'' in the required fields if you wish to remain
anonymous). You can find the petition, status review report, Federal
Register notices, and the list of references electronically on our
website at <a href="https://www.fisheries.noaa.gov/species/queen-conch">https://www.fisheries.noaa.gov/species/queen-conch</a>
FOR FURTHER INFORMATION CONTACT: Calusa Horn, NMFS Southeast Regional
Office, 727-551-5782 or <a href="/cdn-cgi/l/email-protection#7c3f1d10090f1d5234130e123c12131d1d521b130a"><span class="__cf_email__" data-cfemail="397a58554c4a581771564b577957565858175e564f">[email protected]</span></a>, or Maggie Miller, NMFS
Office of Protected Resources, 301-427-8457 or
<a href="/cdn-cgi/l/email-protection#074a66756066756273294f294a6e6b6b6275476968666629606871"><span class="__cf_email__" data-cfemail="3875594a5f594a5d4c167016755154545d4a7856575959165f574e">[email protected]</span></a>.
SUPPLEMENTARY INFORMATION:
Background
On February 27, 2012, we received a petition from WildEarth
Guardians to list the queen conch as threatened or endangered
throughout all or a significant portion of its range under the ESA. We
determined that the petitioned action may be warranted and published a
positive 90-day finding in the Federal Register (77 FR 51763; August
27, 2012). After conducting a status review, we determined that listing
queen conch as threatened or endangered under the ESA was not warranted
and published our determination in the Federal Register (79 FR 65628;
November 5, 2014). In making that determination, we first concluded
that the queen conch was not presently in danger of extinction, nor was
it likely to become so in the foreseeable future. We also evaluated
whether there was a portion of the queen conch's range that was
``significant,'' applying the definition of that term from the joint
U.S. Fish and Wildlife Service/NMFS Policy on Interpretation of the
Phrase ``Significant Portion of Its Range'' (SPR Policy; 79 FR 37580,
July 1, 2014). We concluded that available information did not indicate
any ``portion's contribution to the viability of the species is so
important that, without the members in that portion, the species would
be in danger of extinction, or likely to become so in the foreseeable
future, throughout all of its range.''
WildEarth Guardians and Friends of Animals filed suit on July 27,
2016, in the U.S. District Court for the District of Columbia,
challenging our decision not to list queen conch as threatened or
endangered under the ESA. On August 26, 2019, the court vacated our
determination that listing queen conch under the ESA was not warranted
and remanded the determination back to the NMFS based on our reliance
on the SPR Policy's particular threshold for defining ``significant,''
which was vacated nationwide in 2018 (though other aspects of the
policy remain in effect). See Desert Survivors v. U.S. Dep't of
Interior, 321 F. Supp. 3d 1011 (N.D. Cal. 2018). Following the 2019
ruling of the U.S. District Court for the District of Columbia, we
announced the initiation of a new status review of queen conch and
requested scientific and commercial information from the public (84 FR
66885, December 6, 2019). We received 12 public comments in response to
this request. We also provided notice and requested information from
jurisdictions through the Western Central Atlantic Fishery Commission
(WECAFC), Caribbean Regional Fisheries Mechanism (CRFM), and the
Convention on the International Trade in Endangered Species of Wild
Fauna and Flora (CITES) Authorities. All relevant, new information was
incorporated as appropriate in the status review report and in this
proposed rule. In particular, new information considered in the status
review report includes: (1) fisheries landings data (1950-2018) from
the Food and Agriculture Organization (FAO); (2) reconstructed landing
histories (1950-2016) from the Sea Around Us (SAU) project; (3) results
from recent genetic studies; and (4) the results from regional
hydrodynamics and population connectivity modeling.
Listing Determinations Under the ESA
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 first consider whether a group of organisms
constitutes a ``species'' under section 3 of the ESA, then whether the
status of the species qualifies it for listing as either threatened or
endangered. Section 3 of the ESA defines 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.'' Because the queen conch is an invertebrate, we do not
have the authority to list individual populations as distinct
population segments.
[[Page 55201]]
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.'' Thus, in the
context of the ESA, the Services interpret an ``endangered species'' to
be one that is presently at risk of extinction. A ``threatened
species,'' on the other hand, is not currently at risk of extinction,
but is likely to become so in the foreseeable future. In other words, a
key statutory difference between a threatened and endangered species is
the timing of when a species may be in danger of extinction, either now
(endangered) or in the foreseeable future (threatened). Additionally,
as the definition of ``endangered species'' and ``threatened species''
makes clear, the determination of extinction risk can be based on
either the range-wide status of the species, or the status of the
species in a ``significant portion of its range.'' A species may be
endangered or threatened throughout all of its range or a species may
be endangered or threatened within a significant portion of its range
(SPR).
Section 4(a)(1) of the ESA requires us to determine whether any
species is endangered or threatened as a result of any of the following
five factors: (A) The present or threatened destruction, modification,
or curtailment of its habitat or range; (B) overutilization for
commercial, recreational, scientific, or educational purposes; (C)
disease or predation; (D) the inadequacy of existing regulatory
mechanisms; or (E) other natural or manmade factors affecting its
continued existence (section 4(a)(1)(A)-(E)). Section 4(b)(1)(A) of the
ESA requires us to make listing determinations based solely on the best
scientific and commercial data available after conducting a review of
the status of the species and after taking into account conservation
efforts being made by any State or foreign nation or political
subdivision thereof to protect the species.
Status Review
We convened a team of seven agency scientists to conduct a new
status review for the queen conch and prepare a report. The status
review team (SRT) was comprised of natural resource management
specialists and fishery biologists from the NMFS Southeast Regional
Office, West Coast Regional Office, Office of Protected Resources, and
Southeast Fisheries Science Center (SEFSC). The SRT had group expertise
in queen conch life history and ecology, population dynamics,
connectivity modeling, fisheries management and stock assessment
science, and protected species management and conservation. The status
review report presents the SRT's professional judgment of the
extinction risk facing the queen conch but makes no recommendation as
to the listing status of the species. The status review report was
subjected to independent peer review as required by the Office of
Management and Budget Final Information Quality Bulletin for Peer
Review (M-05-03; December 16, 2004). The status review report was peer
reviewed by three independent specialists selected from the scientific
community, with expertise in queen conch biology and ecology,
conservation and management, and specific knowledge of threats to queen
conch. The peer reviewers were asked to evaluate the adequacy,
appropriateness, and application of data used in the status review as
well as the findings resulting from that data. All peer reviewer
comments were addressed prior to finalizing the status review report.
We subsequently reviewed the status review report, its cited
references, and public and peer reviewer comments. We determined the
status review report, upon which this proposed rule is based, provides
the best available scientific and commercial information on the queen
conch. Much of the information discussed below on queen conch biology
and ecology, distribution and connectivity, density and abundance,
threats, and extinction risk is taken from the status review report.
However, we have independently applied the statutory provisions of the
ESA, including evaluation of the factors set forth in section
4(a)(1)(A)-(E), our regulations regarding listing determinations,
conservation efforts, and the aspects of our SPR Policy that remain
valid in making our determination that the queen conch meets the
definition of a threatened species under the ESA.
Life History, Ecology, and Status of the Petitioned Species
Taxonomy and Species Description
Aliger gigas, originally known as Strombus gigas or more recently
as Lobatus gigas, is commonly known as the queen conch. The queen conch
belongs to the family Strombidae and the most recent classification
places the queen conch under the genus Aliger (Maxwell et al. 2020) in
the class Gastropoda, order Neotaenioglossa, and family Strombidae.
Other accepted synonyms include: Strombus gigas (Linnaeus, 1758);
Lobatus gigas (Linnaeus, 1758); Strombus lucifer (Linnaeus, 1758);
Eustrombus gigas (Linnaeus, 1758); Pyramea lucifer (Linnaeus, 1758);
Strombus samba (Clench 1937); Strombus. horridus (Smith 1940); Strombus
verrilli (McGinty 1946); Strombus canaliculatus (Burry 1949); and
Strombus pahayokee (Petuch 1994), as cited in (Landau et al. 2009).
The queen conch is a large marine gastropod mollusk. Adult queen
conch have a heavy shell (5 pounds, 2.3 kilograms (kg)) with spines on
each whorl of the spire and flared aperture. The shell grows as the
mollusk grows, forming into a spiral shape with a glossy pink interior.
The outside of the shell becomes covered by an organic periostracum
(``around the shell'') layer as the queen conch matures that can be
much darker than the natural color of the shell. Characteristics used
to distinguish queen conch from other family members include: (1)
large, heavy shell; (2) short, sharp spires; (3) brown and horny
operculum; and (4) pink interior of the shell (Prada et al. 2009).
Distribution, Movements, and Habitat Use
The queen conch is distributed throughout the Caribbean Sea, the
Gulf of Mexico, and around Bermuda. Its range includes the following
countries, territories, and areas: Anguilla, Antigua and Barbuda,
Aruba, Barbados, The Bahamas, Belize, Bermuda, Bonaire, British Virgin
Islands, Brazil, Cayman Islands, Colombia, Costa Rica, Cuba,
Cura[ccedil]ao, Dominican Republic, Grenada, Guadeloupe and Martinique,
Guatemala, Haiti, Honduras, Jamaica, Mexico, Montserrat, Nicaragua,
Panama, Puerto Rico, Saba, St. Barthelemy, St. Martin, St. Eustatius,
St. Kitts and Nevis, St. Lucia, St. Vincent and the Grenadines,
Trinidad and Tobago, Turks and Caicos, U.S. Virgin Islands, the United
States (Florida), and Venezuela (Theile 2001; see File S1 in Horn et
al. 2022).
As conch develop they use different habitat types including
seagrass beds, sand flats, algal beds, and rubble areas from a few
centimeters deep to approximately 30 meters (m) (Brownell and Stevely
1981). After the eggs of queen conch hatch, the veligers (larvae) drift
in the water column for up to 30 days depending on phytoplankton
concentration, temperature, and the proximity of settlement habitat.
The minimum pelagic duration is reported from four field studies to be
16 days (Brownell 1977; Davis 1994, 1996;
[[Page 55202]]
Salley 1986), but can range from 21 days to 30 days (Brownell 1977;
D'Asaro 1965; Davis 1994; Paris et al. 2008; Salley 1986) with a mean
of approximately 25 days. These veligers are found primarily in the
upper few meters of the water column (Paris et al. 2008; Posada and
Appeldoorn 1994; Stoner 2003; Stoner and Davis 1997) where they feed on
phytoplankton. When the veligers are morphologically and
physiologically ready, they metamorphose into benthic animals in
response to trophic cues from their seagrass habitat (Davis 2005). The
key trophic cues shown to induce metamorphosis are epiphytes associated
with macroalgae and sediment (Davis and Stoner 1994). Settlement
locations are usually areas that have sufficient tidal circulation and
high macroalgae production. Upon metamorphosis, veligers settle to the
bottom and bury completely into the sediment where they spend much of
their first year of life. They emerge about a year later as juveniles
measuring around 60 millimeters (mm) shell length (Stoner 1989b). When
juvenile conch first emerge from the sediment and move to nearby
seagrass beds, densities can be as high as 200-2000 conch/hectare
(Stoner 1989a; Stoner and Lally 1994; Stoner 2003). A hectare (ha) is
an area 100 meters by 100 meters, equivalent to 2.471 acres.
Queen conch nursery areas primarily occur in back reef areas (i.e.,
shallow sheltered areas, lagoons, behind emergent reefs or cays) of
medium seagrass density, at depths between 2 to 4 m, with strong tidal
currents of at least 50 centimeters (cm)/second (Stoner 1989a), and
frequent tidal water exchanges (Stoner et al. 1996; Stoner and Waite
1991). Seagrass is thought to provide both nutrition and protection
from predators (Ray and Stoner 1995; Stoner and Davis 2010). The
structure of the seagrass beds decreases the risk of predation (Ray and
Stoner 1995), which is very high for juveniles (Appeldoorn 1988c;
Stoner and Glazer 1998; Stoner et al. 2019). Posada et al. (1997)
observed that the most productive nurseries for queen conch tend to
occur in shallow (< 5-6 m deep) seagrass meadows. Jones and Stoner
(1997) found that optimal nursery habitat occurred in areas of medium
density seagrass, particularly areas associated with strong ocean
currents or hydrographic conditions. Boman et al. (2019) observed a
significantly higher probability of positive growth in juvenile conch
in native seagrass compared to invasive seagrass. In The Bahamas,
juveniles were found only in areas within 5 km from the Exuma Sound
inlet, emphasizing the importance of currents and frequent tidal water
exchange that affects both larval supply and growth of their algal food
(Jones and Stoner, 1997). However, there are certain exceptions, such
as in Florida, where many juveniles are found on shallow algal flats,
or in Jamaica, where they can be found on deep banks such as Pedro
Bank.
While the early life stages of queen conch primarily occur in
shallow waters with dense seagrass meadows, adult queen conch can be
found in a wider range of environments (Stoner et al. 1994), including
sand, algal flats, or coral rubble (Acosta 2001; Stoner and Davis
2010). Queen conch are rarely, if ever, found on soft bottoms composed
of silt or mud, or in areas with high coral cover (Acosta 2006). The
movements of adult queen conch are associated with factors like changes
in temperature, food availability, and predation. Adult conch are
typically found in shallow, clear water of oceanic or near-oceanic
salinities at depths generally less than 75 m, but are most common in
waters less than 30 m (McCarthy 2007). Depth limitation is based mostly
on light attenuation limiting their photosynthetic food source (e.g.,
filamentous alga) (McCarthy 2007; Creswell, 1994; Ray and Stoner 1994;
Randall 1964). The average home range size for an individual queen
conch is variable and has been measured at 5.98 ha in Florida (Glazer
et al. 2003), 0.6 to 1.2 ha in Barbados (Phillips et al. 2010), and
0.15 to 0.5 ha in the Turks and Caicos Islands (Hesse 1979). Studies
have suggested that adult conch move to different habitat types during
their reproductive season, but afterwards return to feeding grounds
(Glazer et al. 2003; Stoner and Sandt 1992; Hesse 1979). In general,
adult conch do not move very far from their feeding grounds during
their reproductive season (Stoner and Sandt 1992).
Diet and Feeding
Queen conch are herbivores and primarily feed on macroalgae and
seagrass detritus (Ray and Stoner 1995; Creswell 1994). The production
of red and green algae, which can be highly variable, has been shown to
directly affect the growth of juvenile conch (Stoner 2003; Stoner et
al. 1995; Stoner et al. 1994). Organic material in the sediment
(benthic diatoms and particulate organic matter and cyanobacteria) has
also been suggested to be a source of nutrition to juvenile conch
(Boman et al. 2019; Serviere-Zaragoza et al. 2009; Stoner et al. 1995;
Stoner and Waite 1991). Stoner and Waite (1991) also showed that
macroalgae were the most likely food source of juvenile conch (shell
length 120-140 mm) in native seagrass beds in The Bahamas. Several
studies have indicated that seagrass detritus is an important secondary
food source for juvenile queen conch, in particular detritus of T.
testudinum (Stoner and Waite 1991; Stoner 1989a). In sand habitats,
juveniles can also feed on diatoms and cyanobacteria that are found in
the benthos (Creswell 1994; Ray and Stoner 1995).
Age and Growth
Queen conch are estimated to have a life span of 25-30 years (Davis
2005; McCarthy 2007). As with many gastropods, growth in queen conch is
determinate and strongly influenced by the environment (Mart[iacute]n-
Mora et al. 1995; Alcolado, 1976). The species has determinate growth
and reaches maximum shell length before sexual maturation; thereafter
the shell grows only in thickness (Stoner et al. 2012; Appeldoorn
1988a). Conch are often considered to be mature when the lip is flared,
however Appeldoorn (1988c) observed that the verge (the male
reproductive organ) of thin-lipped males in Puerto Rico was not yet
functional, and true reproductive maturity did not occur until at least
two months after the lip flared outward at about 3.6 years of age. The
result is that thin-lipped individuals probably do not mate or spawn in
the first reproductive season after the shell lip flares, and are at
least 4 years old before first mating. Once the shell lip is formed,
the shell does not increase in length (Appeldoorn 1996; Tewfik et al.
1998). Because the shell lip continues to thicken upon the onset of
maturity (Appeldoorn 1988a), studies have found that shell lip
thickness is a better indicator of sexual maturity rather than the
formation of the flared lip (Appeldoorn 1994b; Clerveaux et al. 2005;
Stoner et al. 2012c). With the onset of sexual maturity, tissue growth
decreases and switches from primarily thickening of the meat to
increasing the weight of the gonads. Once the conch is around ten years
of age, the shell volume starts to decrease, as layers of the shell
mantle are laid down from the inside (Randall 1964). Eventually, the
room inside the shell can no longer accommodate the tissue and the
conch will start to decrease its tissue weight (CFMC and CFRAMP 1999).
Stoner et al. (2012c) found that after shell lip thickness reached 22
to 25 mm, both soft tissue and gonad weight decreased.
[[Page 55203]]
Reproductive Biology
Queen conch reproduce via internal fertilization. Males and females
are distinguished by either a verge (the male reproductive organ) or
egg groove. Approximately three weeks after copulation the female lays
a demersal egg mass on coarse sand of low organic content, completing
deposition within 24-36 hours (D'Asaro 1965; Randall 1964). The egg
mass consists of a long, continuous, egg-filled tube that folds and
sticks together in a compact crescent shape, adhering to sand grains
that provide camouflage and discourage predation. After an incubation
period of approximately five days, the larvae emerge and assume a
pelagic lifestyle (Weil and Laughlin 1984; D'Asaro 1965).
Assessments of fecundity require knowledge of the population sex
ratio, spawning season duration, rate of spawning during the season,
number of eggs per egg mass, and the relationship between body mass and
age (Appeldoorn 1988c). Few studies have investigated these factors
concurrently, and the variability reported in these metrics is high.
For example, estimates of the number of eggs contained within each egg
mass range from 150,000 to 1,649,000 (Appeldoorn 2020; Delgado and
Glazer 2020; Appeldoorn 1993; Berg Jr. and Olsen 1989; Mianmanus 1988;
Weil and Laughlin 1984; D'Asaro 1965; Randall 1964; Robertson 1959).
Additionally, females are capable of storing eggs for several weeks
before laying an egg mass, which means it is possible that multiple
males have fertilized the same eggs (Medley 2008). The ability to store
sperm is advantageous for conch populations since females are still
capable of laying egg masses without encountering another male. The
number of egg masses produced per female is also highly variable and
ranges between 1 and 25 per female per season for experiments performed
in different areas throughout the queen conch range (Appeldoorn 1993;
Berg Jr. and Olsen 1989; Davis et al. 1984; Weil and Laughlin 1984;
Davis and Hesse 1983).
The number of masses produced as well as the number of eggs per
mass may decrease toward the end of the reproductive season (Weil and
Laughlin 1984), but individual variability may also be influenced by
spawning frequency and the size and number of egg masses produced
during the season (Appeldoorn 2020). Differences in spawning rates have
been attributed to spawning site selection, population densities, and
food selection and availability, among other variables. Variability in
spawning activity may also be correlated to water temperature and
weather conditions. For example, reproductive activity decreased with
increasing water turbulence (Davis et al. 1984) and reproduction peaked
with longer days, warmer water temperatures, and relatively stable
circulation patterns (Stoner et al. 1992).
Seasonal movements, usually associated with the initiation of the
reproductive season, are widely known for queen conch. Weil and
Laughlin (1984) reported that adult conch at Los Roques, Venezuela,
moved from offshore feeding areas in the winter to summer spawning
grounds in shallow, inshore sand habitats. In the Turks and Caicos,
adult conch moved from seagrass to sand-algal flats with the onset of
winter (Hesse 1979). Movements to shallower habitats have also been
reported for deep-water populations at St. Croix, U.S. Virgin Islands
(Coulston et al. 1987). Increasing water temperature and photoperiod
are thought to trigger large-scale migrations and the subsequent
initiation of mating. In locations where adult conch are abundant,
these migrations culminate in the formation of reproductive
aggregations. These aggregations generally form in the same locations
each year (Marshak et al. 2006; Glazer and Kidney 2004; Posada et al.
1997) and are dominated by older individuals that produce viable egg
masses (Berg Jr. et al. 1992). However, in some areas large-scale
movements do not occur. For example, in the United States (Florida
Keys), adult aggregations are relatively persistent throughout the
year, although reproductive activity does not occur year-round (Glazer
and Kidney 2004; Glazer et al. 2003). Queen conch found in the deep
waters near Puerto Rico are geographically isolated from nearshore,
shallow habitats and remain offshore during the spawning season
(Garc[iacute]a-Sais et al. 2012). The distribution of feeding and
spawning habitats may also be an important factor in the timing and
extent of adult movements.
Multiple studies involving visual surveys of mating and spawning
events and histological examinations of gonadic activity show that the
duration and intensity of the spawning season varies extensively
throughout the queen conch's range (Table 1 in Horn et al. 2022).
External variables such as temperature, photoperiod, and weather events
interact to mediate seasonality in reproductive and spawning behaviors.
Generally, reproductive activity begins earlier and extends later into
the year with decreasing latitude. Visual surveys of reproductive
activity have reported the reproductive season to extend from May to
September in Florida (D'Asaro 1965), May to November in Puerto Rico
(Appeldoorn 1985), March to September in the Turks and Caicos (Davis et
al. 1984; Hesse 1976), and February through November in the U.S. Virgin
Islands (Coulston et al. 1987; Randall 1964). In warmer regions such as
Cuba and Mexico's Banco Chinchorro, reproductive activity can occur
throughout the year (Cala et al. 2013; Corral and Ogawa 1987; Cruz S.
1986); however, there is a seasonal peak in activity in most areas
during the warmest months, usually from July to September (Aldana-
Aranda et al. 2014).
Spawning Density
Depensatory mechanisms have been implicated as a major factor
limiting the recovery of depleted queen conch populations (Stoner et
al. 2012c; Appeldoorn 1995). Depensation occurs when a population's
decreased abundance or density leads to a reduced per capita growth
rate, thereby reducing the population's ability to recover.
Reproductive potential is primarily reduced by the removal of mature
adults from the population (Appeldoorn 1995). Empirical observations
have suggested mating and egg-laying in queen conch are directly
related to the density of mature adults (Stoner et al. 2012c; Stoner et
al. 2011; Stoner and Ray-Culp 2000). In animals that aggregate to
reproduce, low population densities can make it difficult or impossible
to find a mate (Stoner and Ray-Culp 2000; Erisman et al. 2017; Rossetto
et al. 2015; Stephens et al. 1999; Appeldoorn 1995). Challenges
associated with mate finding are likely exacerbated for slow-moving
animals such as the queen conch (Doerr and Hill 2013; Glazer et al.
2003). This limitation directly impacts the species' ability to
increase its population size because increased ``search time'' depletes
energy resources, reducing the rate of gametogenesis and the overall
reproductive potential of the population. Simulations by Farmer and
Doerr (in review) confirm that limitations on mate finding associated
with density are the primary driver behind observed patterns in queen
conch mating and spawning activity, but similar to field observations
by Gascoigne and Lipcius (2004), it is unlikely to be the only
explanation for lack of reproductive activity at low densities.
An additional postulated depensatory mechanism is the breakdown of
a positive feedback loop between contact with males and the rate of
gametogenesis and spawning in females, where copulation stimulates
oocyte development and maturation, leading to
[[Page 55204]]
more frequent spawning (Appeldoorn 1995). Copulation in conch is more
likely in spawning than non-spawning females, providing an additional
positive feedback mechanism that amplifies the effect at high densities
(Appeldoorn 1988a). Evidence supporting this idea has been provided by
several studies that reported a consistent lag at the start of the
reproductive season between first observations of copulation and first
spawning (Weil and Laughlin 1984; Brownell 1977; Hesse 1976; Randall
1964). This lag period, averaging three weeks, may represent the time
required to achieve oocyte maturation after first copulation. Farmer
and Doerr (in review) considered differences in adult density, movement
speeds, scent-tracking, barriers to movement, interbreeding rest
periods, perception distance, and sexual facilitation. Sexual
facilitation was the only mechanism explaining the lack of empirical
observations of mating at relatively low population densities,
providing statistical confirmation that the reductions of densities
caused by overfishing of spawning aggregations increases the
probability of recruitment failure beyond what would be anticipated
from delays in mate finding alone. This is consistent with observations
by Gascoigne and Lipcius (2004), which indicate that in addition to
depensatory mechanisms associated with mate finding, delayed functional
maturity at low density sites can explain declines in reproductive
activity.
Because direct physical contact is necessary for copulation and
queen conch are slow moving, the density of mature adults within
localized queen conch populations is a critical and complex factor
governing mating success and population sustainability. Although many
surveys of conch populations have been completed over the last half
century, few studies have simultaneously investigated the relationship
between adult density and reproductive rates. Of these, the reported
rates of reproductive activity associated with surveys of adult
populations have varied extensively across multiple jurisdiction as
density is dependent on the scale of measurement and the targeted area
surveyed. For example, in The Bahamas where queen conch populations are
at densities near 200 adults per hectare, Stoner and Ray-Culp (2000)
reported mating and spawning rates of approximately 13 percent and 10
percent, respectively. During continued surveys in fished areas (Berry
and Andros Islands) and a no-take reserve (Exuma Cays Land and Sea
Park) of The Bahamas, Stoner et al. (2012c) observed that, at a mean
adult density of 60 conch/ha within the Exuma Cays Land and Sea Park,
9.8 percent of adult queen conch were mating, while at 118 adult conch/
ha at Andros Island, approximately 2.4 percent were mating, and at 131
adult conch/ha at the Berry Islands, only 5.9 percent were involved in
mating activity. Doerr and Hill (2018) reported reproductive activity
in 2.4 percent of adult conch located across the shelf of St. Croix,
U.S. Virgin Islands, with the lowest mean density of adult queen conch
at survey sites, where reproductive activity occurred, was 63.7 adult
conch/ha. Of these studies, the highest densities were reported from
Cuba, where at one protected site with densities of 223 adult conch/ha
only 0.3 percent of adult queen conch were mating, while at another
site with a reported adult density of 497 conch/ha, 3.7 percent of
conch were mating, and 2.5 percent were involved in spawning (Cala et
al. 2013). In Colombia, however, reproductive activity demonstrated by
the presence of egg masses was reported in areas with population
densities as low as 24 and 11 conch/ha (G[oacute]mez-Campo et al.
2010). The scale over which these observations were recorded and
subsequent interpretation of the spatial dispersion of queen conch are
critical to understanding differences among study conclusions.
As previously discussed, queen conch life history traits make them
vulnerable to depensatory mechanisms. When reproductive fitness
declines such that per capita population growth rate becomes negative,
localized extinction may result (Courchamp et al. 1999; Allee 1931).
Appeldoorn (1988a) initially suggested that queen conch may have a
critical density for egg production, and Stoner and Ray-Culp (2000)
provided evidence for demographic effects in queen conch populations,
reporting a complete absence of mating and spawning in population
densities less than 56 and 48 adult conch/ha, respectively. They
concluded that the absence of reproduction in low-density populations
was primarily related to encounter rate and noted that reproductive
activity reached an asymptotic level near 200 adult conch/ha (Stoner
and Ray-Culp 2000). Based on these studies, 50 adult conch/ha is
generally accepted as the minimum threshold required to achieve some
level of reproductive activity within a given conch population
(Gascoigne and Lipcius 2004; Stoner and Ray-Culp 2000; Stephens and
Sutherland 1999; Appeldoorn 1995). Conversely, Delgado and Glazer
(2020) reported the highest adult queen conch threshold densities below
which no reproduction was observed, with no mating occurring at
aggregation densities below 204 adult conch/ha and no spawning at
aggregation densities below 90 adult conch/ha. Given the highly
aggregated nature of queen conch (Glazer and Kidney 2004; Glazer et al.
2003), managing for minimum cross-shelf densities (i.e., 100 adult
conch/ha) does not specifically protect the high-density spawning
aggregations where most reproduction occurs. Thus, the Delgado and
Glazer (2020) contend that queen conch fishery managers should identify
and protect high density queen conch spawning aggregations irrespective
of cross-shelf densities.
The persistent formation of adult queen conch aggregations may help
to sustain some populations as evidenced by long-term intra-aggregation
surveys conducted by Delgado and Glazer (2020) in Florida, which show
that, as aggregation densities increase both mating and spawning
increase, correspondingly. Delgado and Glazer (2020) observed an
increase in mating activity, peaking at 71 percent of the aggregation
at densities greater than 800 adult conch/ha. In addition, a greater
portion of the aggregations were found to have egg-laying females as
aggregation density increased. The percentage of aggregations with
spawning females reached a peak of just over 84 percent at aggregation
densities greater than 600 adult conch/ha (Delgado and Glazer 2020).
Similarly, Stoner et al. (2012b) reported that mating frequency
increased at higher densities of adults in The Bahamas, with a maximum
of 34 percent of the population mating at approximately 2,500 adult
conch/ha. Repeat visual surveys in the same sites in The Bahamas have
provided evidence of this susceptibility, revealing that adult
densities in the Exuma Cays Land and Sea Park have declined
significantly over 22 years due to lack of recruitment (Stoner et al.
2019). Stoner et al. (2019) further concluded that most conch
populations in The Bahamas are currently at or below critical densities
for successful mating and reproduction and that significant management
measures are needed to preserve the stock. Similar long-term declines
of reproductively active adult conch have been reported within the Port
Honduras Marine Reserve in southern Belize. Densities of conch in the
Port Honduras Marine Reserve (no-take zone) have been declining since
2009, falling below
[[Page 55205]]
88 conch/ha by 2013, decreasing further to fewer than 56 adult conch/ha
in 2014 (Foley 2016, unpublished. cited in, Foley and Takahashi 2017).
If queen conch, particularly females, do not have the opportunity to
mate and spawn to their full potential, fewer offspring are produced
per individual, which is likely to lead to a decrease in the per capita
population growth rate (Gascoigne et al. 2009). Therefore this is a
critical consideration in assessing the sustainability of conch
populations. As discussed above, although the observed minimum
reproductive density thresholds are highly variable, queen conch
populations are recommended to be managed to maintain a threshold
density of 100 adult conch/ha (Prada 2017). A density value of 100
adult conch/ha is recommended as a minimum reference threshold for
successful reproduction, following a recommendation from the Queen
Conch Expert Workshop, held in May 2012 in Miami, Florida (FAO 2012).
The Regional Queen Conch Fisheries Management and Conservation Plan
(Prada 2017) and the United Nations Environment Programme (UNEP) have
both adopted 100 adult conch/ha as the minimum density threshold to
avoid significant impacts to recruitment (UNEP 2012). Unfortunately,
many queen conch populations do not meet the conditions necessary for
successful reproduction and sustainability because adult queen conch
densities in most jurisdictions are below 100 adult conch/ha (see
Status of the Population below).
Population Structure and Genetics
Early studies using allozymes (variant forms of the same enzyme) to
examine the genetic structure of queen conch implied high levels of
gene flow, but also showed isolated genetic structure for populations
either at isolated sites or at the microscale level.
Mitton et al. (1989) collected samples from nine locations across
the Caribbean including Bermuda, Turks and Caicos, St. Kitts (St.
Christopher) and Nevis, St. Lucia, the Grenadines, Bequia Island,
Barbados, and Belize, and reported high gene flow as well as genetic
differentiation at all spatial scales. For example, they found that
queen conch in Bermuda and Barbados were genetically isolated from the
rest of the sampled locations. Yet, they also found that conch sampled
at two geographically close locations (i.e., Gros Inlet and Vieux Fort)
in St. Lucia had significant genetic differentiation despite being
separated by only 40 km (Mitton et al. 1989). Conch sampled in the
United States (Florida Keys) also demonstrated significant spatial and
temporal genetic variation, although genetic similarity among
populations was high (Campton et al. 1992). Tello-Cetina et al. (2005)
sampled conch from four sites along the Yucatan Peninsula and reported
relatively high levels of intrapopulation diversity and little
geographic differentiation, with the population from the Alacranes Reef
having the furthest genetic distance from the other three sites.
Several studies conducted in Jamaica reported similar levels of
connectivity and genetic differentiation. Blythe-Mallett et al. (2021)
sampled multiple zones across Pedro Bank, an important commercial
fishing ground southwest of Jamaica, and identified two possible
subpopulations, one on the heavily exploited eastern end of the bank
and another on the central and western end. Pedro Bank is directly
impacted by the westward flow of the Caribbean current and could serve
as the primary recruitment area of queen conch larvae from upstream
locations (Blythe-Mallett et al. 2021). Pedro Bank is geographically
isolated and receives limited gene flow from mainland Jamaica and other
historically important offshore populations within the Jamaican
Exclusive Economic Zone (EEZ) (Kitson-Walters et al. 2018). The high
degree of genetic relatedness within conch sampled from Pedro Bank
likely indicates that the populations are sufficiently self-sustaining
(Kitson-Walters et al. 2018), but still receive larvae from upstream
sources that contribute to the population on the eastern end of the
bank (Blythe-Mallett et al. 2021).
Studies conducted in the Mexican Caribbean have also detected a
spatial genetic structure for queen conch populations. P[eacute]rez-
Enriquez et al. (2011) identified a genetic cline along the southern
Mexican Caribbean to north of the Yucatan Peninsula, with a reduced
gene flow observed between the two most distant locations, representing
an increase in genetic differences as geographic distance increased.
These authors suggested that since the overall genetic diversity varied
from medium to high values, the queen conch had not reached genetic
level indicative of a population bottleneck (P[eacute]rez-Enriquez et
al. 2011). Machkour-M'Rabet et al. (2017) used updated molecular
markers to analyze queen conch from seven sites within the same area
and observed similar results with the exception of the apparent genetic
isolation of queen conch collected on Isla Cozumel, which was not
detected by P[eacute]rez-Enriquez et al. (2011). The results of this
study led Machkour-M'Rabet et al. (2017) to conclude that populations
of queen conch along the Mesoamerican Reef are not panmictic and
demonstrate genetic patchiness indicative of homogeneity among sample
areas, providing further evidence for the pattern of isolation by
distance.
M[aacute]rquez-Pretel et al. (2013) found four genetic stocks
reflecting heterogeneous spatial mosaics of marine dispersion between
the San Andres archipelago and the Colombian coastal areas. Queen conch
in these areas exhibited an overall deficit of heterozygosity related
to assortative mating or inbreeding, potentially leading to a loss in
genetic variation (M[aacute]rquez-Pretel et al. 2013).
A broad-ranging spatial genetic study of queen conch across the
greater Caribbean using nine microsatellite DNA markers (Truelove et
al. 2017) found that basin-wide gene flow was constrained by oceanic
distance that served to isolate local populations. Truelove et al.
(2017) genetically characterize 643 individuals from 19 locations
including Florida, The Bahamas, Anguilla, the Caribbean Netherlands
(i.e., Bonaire, St Eustatius, and Saba), Jamaica, Honduras, Belize, and
Mexico, and determined that queen conch do not form a single panmictic
population in the greater Caribbean. The authors reported significant
differentiation between and within jurisdictions and among sites
irrespective of geographic location. Gene flow was constrained by
oceanic distance and local populations tended to be genetically
isolated.
Recently, Douglas et al. (2020) conducted a genomic analysis using
single nucleotide polymorphisms from two northeast Caribbean Basin
Islands (Grand Bahama to the north and Eleuthera to the south). The
authors identified distinct populations on the south side of Grand
Bahama Island and the west side of Eleuthera Island potentially due to
larval separation by the Great Bahama Canyon. Despite extensive spatial
separation of sampled populations around Puerto Rico, Beltr[aacute]n
(2019) concluded that there was little genetic structure in the conch
population. However, genetic analyses of four visually characterized
phenotypes showed that one morph (designated as Flin) was slightly
differentiated from the other phenotypes sampled. Further research into
this aspect of queen conch biology is needed to examine the degree of
differentiation between phenotypes and to determine if they share the
same distribution across the Caribbean region. The results presented in
all of these studies provide evidence that variation in marine
currents, surface winds, and
[[Page 55206]]
meteorological events can either promote larval dispersal or act as
barriers enhancing larval retention.
Status of the Population
The SRT reviewed data from 39 jurisdictions throughout the species'
range and developed several interrelated assessments that were used to
inform the status of the queen conch. First, the SRT compiled cross-
shelf adult conch density estimates for each jurisdiction in the
species' range (see Density Estimates below). Second, the SRT developed
spatially explicit habitat estimates (see Conch Habitat Estimate below)
for each jurisdiction. The habitat estimates were necessary for the SRT
to be able to estimate total abundance and evaluate population
connectivity. Third, the SRT extrapolated each jurisdiction's conch
density estimate in the surveyed areas to the jurisdiction's total
estimated habitat area to generate population abundance estimates at a
jurisdiction-level (see Abundance Estimates below). Last, the SRT
evaluated population connectivity to elucidate the potential impacts of
localized low conch densities on population-wide connectivity patterns
(see Population Connectivity below). As described above, queen conch
reproductive failure has been attributed in many cases to declines in
population densities. There are two density thresholds (i.e., <50 adult
conch/ha and >100 adult conch/ha) that are well established in the
scientific literature and are generally accepted by fisheries managers.
The scientific literature indicates that when adult queen conch numbers
decline to fewer than 50 adult conch/ha there are significant
implications for finding a mate and thus reproductive activity and
population growth. When adult queen conch density are reduced to this
degree, reproductive activity is limited or non-existent. Along those
same lines, the available literature suggests that populations with
adult queen conch densities greater than 100 adult conch/ha are
sufficient in most cases to promote successful mate finding and thus
reproductive activity and population growth. The 100 adult conch/ha
density threshold recommendation was prepared by the Queen Conch Expert
Working Group (Miami, Florida, May 2012), and subsequently accepted by
consensus by fisheries managers participating in the WECAFC/Caribbean
Fishery Management Council (CFMC)/Organization of the Fisheries and
Aquaculture Sector of the Central American (OSPESCA)/CRFM Working
Group, as minimum reference point or ``precautionary principle''
required to sustain conch populations (Prada et al. 2017).
Considering this information, including the best available
scientific and commercial information on queen conch reproduction,
depensatory processes, and population growth, the SRT applied the
following density thresholds to queen conch populations:
<bullet> Populations with densities below the 50 adult conch/ha
threshold are considered to be not reproductively active due to low
adult encounter rates or mate finding. This threshold is largely
recognized as an absolute minimum required to support mate finding and
thus reproduction.
<bullet> Populations with densities between 50-99 adult conch/ha
are considered to have reduced reproductive activity resulting in
minimal population growth.
<bullet> Populations with densities above 100 adult conch/ha are
considered to be at a density that supports reproductive activity
resulting in population growth.
These density thresholds were used to evaluate the status of queen
conch populations in each jurisdiction, and to assess how heterogeneous
fishing pressure and localized depletion (i.e., low adult queen conch
densities, leading to reduced egg and larval production) effect
population connectivity throughout the species' range. The results of
these assessments are described in the following sections.
Density Estimates
In order to develop estimates of queen conch density, the SRT
conducted a comprehensive, jurisdiction-by-jurisdiction search to
identify literature pertaining to the status of queen conch throughout
its range. The SRT reviewed the best scientific and commercial
information including all relevant published and gray literature,
databases, and reports. The SRT organized this information and data by
jurisdiction and searched systematically for information on queen conch
densities. The SRT also considered relevant information provided during
the public comment period (84 FR 66885, December 6, 2019). The SRT's
goal was to compile robust, cross-shelf adult queen conch density
estimates for each jurisdiction. To the extent possible, the SRT
focused on the most recent studies where randomized sampling was
conducted across broad areas of the shelf, including a range of
habitats and depths. For jurisdictions where such studies were not
available, the SRT used available density information. For example, in
some cases the only available data were single point estimates from a
study or workshop report. For nine jurisdictions where no density
information was available (i.e., Cura[ccedil]ao, Costa Rica, Dominica,
Grenada, Montserrat, St. Kitts and Nevis, St. Martin, St. Barthelemy,
and Trinidad and Tobago), the SRT approximated queen conch density
estimates based on density estimates for the nearest neighboring
jurisdiction that had information available. The SRT used available
qualitative information on the general population status (e.g.,
severely depleted, moderately fished, and lightly exploited) to ensure
that approximating queen conch densities based on a jurisdiction's
nearest neighbor was reasonable (for detailed discussion on methods see
Horn et al. 2022).
From each study or report compiled, the SRT noted the location,
year of the survey (1996 to 2022), total area surveyed, status of the
area surveyed (fished or unfished), and the survey methods used (see
Table 2 in Horn et al. 2022). The SRT extracted information on the
overall density or the adult density (or both) of queen conch, and
recorded these in a spreadsheet and standardized to a per hectare (ha)
unit (see S5 in Horn et al. 2022). For jurisdictions with large shelf
areas (e.g., The Bahamas, Belize, Mexico) densities were recorded at
the sub-jurisdiction level (e.g., as defined by region, bank, or
cardinal direction from an island). For smaller jurisdictions (e.g.,
those within the Lesser Antilles), queen conch densities were typically
reported for an entire island or group of islands. The status review
report (Horn et al. 2022) provides additional detail on how the SRT
estimated queen conch population densities.
The adult queen conch density estimates were also plotted by their
geographical locations (see Figure 6 in Horn et al. 2022). The results
revealed that several jurisdictions, mostly located in the north-
central to the southwestern Caribbean (i.e., Turks and Caicos, The
Bahamas' Cay Sal Bank and Jumentos and Ragged Cays, Cuba, Jamaica,
Nicaragua, Costa Rica), tended to have higher adult conch population
densities (>100 adult conch/ha) indicating that these populations are
reproductively active and are supporting successful population growth.
There are a two jurisdictions (i.e., St. Eustatius and St. Kitts and
Nevis) within the eastern Caribbean region and a single jurisdiction
(i.e., Cayman Islands) in the central Caribbean region, that have
moderate adult conch population densities (<100 adult conch/ha, but >50
adult conch/ha). In the eastern Caribbean only two jurisdictions (St.
Lucia and Saba) have queen conch densities greater than 100 adult
conch/ha. With a few exceptions, the rest of
[[Page 55207]]
the jurisdictions not previously mentioned above (i.e., Aruba,
Anguilla, Antigua and Barbuda, Barbados, Belize, Bermuda, Bonaire, The
Bahamas' Western and Central Great Banks and Little Bahama Bank,
British Virgin Islands, Colombia's Serranilia and Quitasueno Banks,
Cura[ccedil]ao, Dominica, Dominican Republic, Grenada, Guadeloupe,
Haiti, Martinique, Mexico, Montserrat, Panama, Puerto Rico, St,
Barthelemy, St. Martin, St. Vincent and Grenadines, Trinidad and
Tobago, United States (Florida), U.S. Virgin Islands, and Venezuela),
have queen conch densities near or below the minimum adult queen conch
density threshold (<50 adult conch/ha) required to support reproductive
activity. These jurisdictions represent approximately 27 percent
(19,626 km\2\) of the estimated habitat available in the Caribbean
region.
Conch Habitat Estimate
To increase the SRT's understanding of the status of queen conch
throughout its range, the SRT estimated conch habitat and prepared a
spatially explicit map for the Caribbean region. This spatially
explicit conch habitat estimate was necessary in order for the SRT to
estimate total abundance and conduct the population connectivity
analysis. To develop an estimate of habitat area, the SRT conducted an
extensive search for the best available habitat information, including
estimated conch fishing bank areas, and contacted researchers and
institutions involved in various mapping efforts. The SRT determined
that a 0-20 m depth habitat area represented a best estimate because
the available information indicates that conch are found in shallow
waters generally less than 20 m depth (Berg Jr. et al. 1992; Boidron-
Metairon 1992; Delgado and Glazer 2020; Salley 1986; Stoner and Sandt
1992; Stoner and Schwarte 1994). The most comprehensive and suitable
publicly-available habitat map that could be found was the Millennium
Coral Mapping Project, which specifies 1,359 8-km by 8-km polygons
based on coral reefs locations (Andr[eacute]fou[euml]t et al. 2001).
The polygons included seagrass and coral reef locations where queen
conch occur (Kough 2019; Souza Jr. and Kough 2020). To ensure that all
spawning sites, including deep water spawning sites (i.e., at depths
greater than 20 m), were included in the dataset, the SRT verified the
habitat map with spawning sites reported in the available literature
(Berg Jr. et al. 1992; Brownell 1977; Cala et al. 2013; Coulston et al.
1987; D'Asaro 1965; Davis et al. 1984; de Graaf et al. 2014;
Garc[iacute]a E. et al. 1992; Gracia-Escobar et al. 1992; Lagos-Bayona
et al. 1996; M[aacute]rquez-Pretel et al. 1994; Meijer zu Schlochtern
2014; P[eacute]rez-P[eacute]rez and Aldana-Aranda 2003; Randall 1964;
Stoner et al. 1992; Truelove et al. 2017; Weil and Laughlin 1984;
Wicklund et al. 1991; Wilkins et al. 1987; Wynne et al. 2016).
Following this review, the SRT included 13 additional deep spawning
sites for Venezuela, Cuba, The Bahamas, U.S. Virgin Islands, Turks and
Caicos, Saba, Colombia, Belize, Honduras, and Jamaica (Brownell 1977;
Cala et al. 2013; Davis et al. 1984; De Graaf et al. 2014; Lagos-Bayona
et al. 1996; Randall 1964; Stoner et al. 1992; Truelove et al. 2017;
Weil and Laughlin 1984; Wicklund et al. 1991). The SRT also
incorporated 13 shallow polygons not initially present in the dataset
for St. Eustatius, U.S. Virgin Islands, Colombia, United States
(Florida), Mexico, Jamaica, Saba, Bonaire and The Bahamas (Meijer zu
Schlochtern 2014; Randall 1964; Coulston et al. 1987; Gracia-Escobar et
al. 1992; M[aacute]rquez-Pretel et al. 1994, Truelove et al. 2017).
Overall, the habitat area estimates from the data source selected by
the SRT were much lower than total seagrass area estimates, and
generally ranged from approximately 30 to 100 percent of the estimated
conch fishing banks and incorporated known deep-water spawning sites
(see Figure 5 in Horn et al. 2022). Thus, the SRT concluded, and we
agree, that its habitat estimates were likely conservative, but
suitable for analysis of general connectivity patterns and population
abundance estimates.
Abundance Estimates
The SRT estimated abundance by extrapolating adult queen conch
density estimates across the estimated habitat areas. However, the SRT
used these abundance estimates with caution because the available
density estimates on which they are based were dated, had sparse data,
or were conducted in small areas. In some cases, the number of
available surveys with queen conch densities were also limited. For
example, the very high estimated queen conch abundance from Cuba is
particularly questionable due to the small sample size of survey and
the large shelf area over which the survey density data was expanded.
Where no survey data were available (i.e., Costa Rica, Cura[ccedil]ao,
Dominica, Grenada. St. Kitts and Nevis, St. Barthelemy, St. Martin,
Monserrat, and Trinidad and Tobago), density estimates were
approximated from the nearest neighboring jurisdiction, and thus their
abundance estimates are highly uncertain. The estimated conch habitat
areas also introduce some uncertainty in the estimates, and the
resolution of the SRT's habitat map is coarse (for additional
discussion on methods see Horn et al. 2022).
Despite the aforementioned constraints, the SRT estimated
jurisdiction-level conch abundance by multiplying available conch
density estimates by estimated habitat areas. This approach assumed the
range of jurisdiction-level survey-generated conch density estimates is
representative of the range of conch densities across the entirety of
each jurisdiction's estimated habitat area. When available, multiple
surveys were used to better capture the substantial uncertainty
inherent in this approach. In jurisdictions where comprehensive surveys
were carried out across all areas of the shelf, the mean estimates
reported from each survey typically take into account any sub-
jurisdiction level variability in conch densities; however, in cases
where extrapolations were based on only a few reported density
estimates or sampling that was done over a small area, this assumption
may be violated. In most studies, conch densities were surveyed across
various habitat types (including those types supporting few or no
conch) and weighted averages were reported. Thus, those survey means
account for areas of both high and low density. The SRT also made
efforts to quantify the uncertainty inherent in basing the abundance
estimates on surveys that used different methodologies, occurred over a
wide time span and over a range of spatial scales. The results suggest
that adult queen conch abundance is estimated (i.e., the sum of median
estimated abundance across all jurisdictions) to be about 743 million
individuals (90 percent confidence interval of 450 million to 1.492
billion). Adult queen conch abundance was estimated to be between ten
and 100 million individuals in six jurisdictions, and 15 jurisdictions
had median estimated abundances between one and ten million adults. The
estimated adult abundance was less than one million adults in each of
20 jurisdictions, with three of those jurisdictions estimated to have
populations of fewer than 100,000 adult queen conch. Seven
jurisdictions (i.e., Cuba, The Bahamas, Nicaragua, Jamaica, Honduras,
the Turks and Caicos Islands, and Mexico) accounted for 95 percent of
the population of adult queen conch. Within the species' range, Cuba,
The Bahamas, and Nicaragua, are estimated to have the most conch
habitat area (56 percent) and the majority of adult queen conch
[[Page 55208]]
population abundance (84.1 percent). In addition, Jamaica, Honduras,
Turks and Caicos, and Mexico are the other major contributors, in terms
of both habitat area and conch abundance (see Figures 10, 11, in Horn
et al. 2022). Twenty-one jurisdictions make up 95 percent of the total
estimated conch habitat area, while only seven jurisdictions (i.e.,
Cuba, the Bahamas, Nicaragua, Jamaica, Honduras, Turks and Caicos, and
Mexico) make up 95 percent of the total estimated abundance. This
indicates that conch are depleted in many jurisdictions with large
habitat areas, and the remaining populations are concentrated in just a
few jurisdictions (Horn et al. 2022).
Population Connectivity
To elucidate the potential impacts of localized low adult conch
densities on population-wide connectivity patterns, the SRT evaluated
queen conch population connectivity. The population connectivity model
was based on a simulation of the entire pelagic phase of the conch
early life cycle, from the hatching of eggs to the settlement of conch
veligers in suitable habitats (Vaz et al. 2022). This population
connectivity evaluation offers insights into how overall exchange of
larvae across the species' range has been impacted by overexploitation
of adult conch in certain areas. Two sets of simulations were
conducted. First, the connectivity patterns were simulated for uniform
egg releases across the entire Caribbean region (from 8[deg]N to
37[deg]N and from 98[deg]W to 59[deg]W); this represents an
``unexploited spawning'' historical density scenario in which all
jurisdictions have the same potential for reproductive levels, on a
per-area basis. A second simulation of connectivity patterns
representing an ``exploited'' scenario, incorporated realistic
localized density patterns by scaling the number of eggs released (on a
per-area basis, by jurisdiction or region) by the adult conch
densities, and accounts for Allee effects at very low densities (<50
adult conch/ha). Two different hydrodynamic models were used to
simulate larvae dispersal through oceanic processes (e.g., oceanic
circulation, velocities, sea surface temperatures) (For detailed
discussion on methods see Horn et al. 2022).
The comparison of the two sets of simulations illustrates the
population-level impact of heterogeneous patterns in densities of adult
conch (see Figure 12 in Horn et al. 2022). The most apparent
differences in the two sets of simulations emerged from the fact that
many of the jurisdictions had conch densities well below the critical
threshold for reproduction (<50 adult conch/ha) and were considered to
be reproductively non-viable. Within the ``exploited'' scenario, the
SRT assumed no larvae were spawned from these jurisdictions;
subsequently they could only act as sinks (e.g., populations that are
not contributing or receiving larvae) for queen conch larvae to settle,
but were not sources for themselves or other locations. Connectivity
patterns emerging from ``exploited'' scenario were thus drastically
different (see Figure 12 in Horn et al. 2022). For example, due to
their position up current and their small shelf areas, the Lesser
Antilles (i.e., Leeward and Windward Islands) were estimated to be
historically important for contributing larval input to other
jurisdictions downstream (i.e., to the west). However, due to low adult
conch densities in many of these jurisdictions, they are no longer
expected to contribute larvae in the ``exploited'' scenario, resulting
in reduced larval input into the Greater Antilles and Colombia.
Other patterns in comparing the ``unexploited'' versus and
``exploited'' simulations were more subtle, but would be locally
significant. For example, historically the Turks and Caicos Islands
were estimated to have received many larvae from the Dominican Republic
and Haiti, which would have been important given its low local
retention rate (see Figure 12 in Horn et al. 2022). However, due to low
adult conch densities in these source jurisdictions, the ``exploited''
scenario suggests that Turks and Caicos Islands are now entirely
dependent on local production, and a substantial percentage of larvae
are exported to The Bahamas. Likewise, the ``unexploited'' simulation
suggests that the United States (Florida) was dependent on relatively
high local retention, with the most significant external source of
larvae coming from Mexico (see Figure 12, left column in Horn et al.
2022). Both Florida and Mexico are thought to now have very low adult
queen conch densities (<50 conch/ha) unable to support any reproductive
activity; in other words, Florida currently has no significant upstream
or local sources of larvae. This could explain why, despite a
moratorium on fishing for several decades, queen conch in Florida
waters have been slow to recover (Glazer and Delgado 2020).
The SRT also found that some jurisdictions acted as important
``connectors'' between different regions of the population as a whole,
and could be important for maintaining genetic diversity. The
importance of a jurisdiction as a ``connector'' was quantified
mathematically as a Betweenness Centrality (BC) value on a scale of 0
to 1. The BC value measures the relative influence of a jurisdiction's
conch reproductive output on the flow of larvae (e.g., larvae dispersed
and retained) among jurisdictions range wide. The median of all
calculated BC values (approximately 0.05-0.06) was selected to
distinguish between high versus low BC values (Vaz et al. 2022), which
is appropriate given that the BC values are a relative scale of non-
normally distributed values. Jurisdictions with high BC values (above
the median) act as ecological corridors that facilitate larval flow and
are essential to preserve population connectivity. The ``unexploited''
scenario identified Jamaica, Cuba, and the Dominican Republic as having
a high BC value, and to a lesser extent Puerto Rico and Colombia (see
Figure 13 in Horn et al. 2022). This was not surprising given the
relative central location of these jurisdictions and the exposure of
their shelves to a diversity of ocean currents, which allows them to be
``connectors'' of larval flow. In contrast, jurisdictions located at
the most up current (e.g., Lesser Antilles) or down current locations
(e.g., Florida, Bermuda), or those located at the fringes of the region
(e.g., Panama, Bermuda) were not identified as important connectors of
larval flow and, as expected, had low BC values (below the median) (see
Figure 13 in Horn et al. 2022).
Jurisdictions with documented low adult conch densities influenced
the estimated connections between jurisdictions when comparing the
``unexploited'' to ``exploited'' scenarios. One of the biggest
differences was the absence of reproductive output (e.g., larval
recruits) from Puerto Rico, Dominican Republic, and Haiti. These
jurisdictions had a high BC value (i.e., above 0.05-0.06) under the
``unexploited'' scenario, but have a low BC value (i.e., below 0.05)
under the ``exploited'' scenario because they no longer function as
important connectors (see Figure 13a in Horn et al. 2022). An almost
complete break in the connectivity between the eastern and western
Caribbean region was apparent in the ``exploited'' scenario, with the
Dominican Republic receiving limited larvae from Cuba, Turks and
Caicos, and from a deep mesophotic reef off the west coast of Puerto
Rico. When those jurisdictions were removed from the chain of larval
supply in the ``exploited'' scenario, Jamaica and Cuba remained
important connectors in the
[[Page 55209]]
western portion of the range, and some of the offshore banks in
Colombia remained functional connectors (see Figure 13 in Horn et al.
2022). While Vaz et al. (2022) indicates that connections have been
lost in several locations due to the existence of low adult conch
densities, points of connection likely still exist, albeit reduced,
which allow some exchange of larvae and maintenance of some genetic
diversity.
Localized patterns of conch overfishing can also influence
genetics. The SRT estimated genetic distance between jurisdictions and
then compared those to a Caribbean-wide genetic study (Vaz et al. 2022;
Truelove et al. 2017). The ``unexploited'' scenario corresponded well
to the patterns observed by Truelove et al. (2017) given that larvae
within each region identified by the Truelove et al. (2017) were most
likely locally originated. The exception was the high probability of
larval exchange between The Bahamas and Turks and Caicos Islands and
the Greater Antilles (see Figure 12 in Horn et al. 2022). In the
``exploited'' scenario, six of the 12 jurisdictions sampled by Truelove
et al. (2017) were not reproductively active (Vaz et al. 2022). Due to
the lack of spawning, it was expected that not all connectivity
patterns could be reproduced. Indeed, in this case, the high self-
settlement observed for Mexico, Belize, and Florida was absent due to
the lack of reproductive activity (Vaz et al. 2022). Subsequently, the
genetic evaluation focused only on the results of the ``unexploited''
scenario since the results of the ``exploited'' scenario were
insignificant due to the reduced number of data points (i.e.,
jurisdictions). The results suggest that queen conch populations
exhibit an isolation-by-distance pattern (Vaz et al. 2022).
Summary of Factors Affecting Queen Conch
As described above, section 4(a)(1) of the ESA and NMFS'
implementing regulations (50 CFR 424.11(c)) state that we must
determine whether a species is endangered or threatened because of any
one or a combination of the following factors: (A) The present or
threatened destruction, modification, or curtailment of its habitat or
range; (B) overutilization for commercial, recreational, scientific, or
educational purposes; (C) disease or predation; (D) inadequacy of
existing regulatory mechanisms; or (E) other natural or manmade factors
affecting its continued existence. The SRT summarized information
regarding each of these threats according to the factors specified in
section 4(a)(1) of the ESA. We conclude the SRT's findings with respect
to the ESA section 4(a)(1) listing factors are well-considered and
based on the best available scientific information, and we concur with
their assessment. Available information does not indicate that
destruction, modification or curtailment of the species' habitat or
range and disease or predation are operative threats on this species;
therefore, we do not discuss those further here. More details with
respect to the available information on these topics can be found in
the status review report (Horn et al. 2022). This section briefly
summarizes the SRT's findings regarding the following factors:
overutilization for commercial, recreational, scientific, or
educational purposes, inadequacy of existing regulatory mechanisms; and
other natural or manmade factors affecting its continued existence.
Overutilization for Commercial, Recreational, Scientific, or
Educational Purposes
Description of the Fishery
Queen conch have been harvested for centuries and are an important
fishery resource for many nations in the Caribbean and Central America.
The most common product in trade is queen conch meat. The FAO landings
data indicate that the total annual landings in 2018 (most recent year
data are available) for all jurisdictions is estimated to be 33,797
metric tons (mt) (see S1; Horn et al. 2022). Prada et al. (2017)
estimated production of queen conch meat for most jurisdictions is
approximately 7,800 mt annually. However, total conch production is
difficult to estimate because of incomplete and incomparable data
across jurisdictions (Prada et al. 2017). The majority of the queen
conch meat is landed in Belize, The Bahamas, Honduras, Jamaica,
Nicaragua, and Turks and Caicos. In the artisanal fishery, queen conch
are sometimes landed with the shell, but mostly as unclean meat with
the majority of organs still attached. Additionally, local markets and
subsistence fishing of queen conch is often not monitored or not
included in catch data. In some jurisdictions, the subsistence and
locally marketed catches are small, but they can be high in some
jurisdictions (Prada et al. 2017). Furthermore, the best estimates of
unreported catch and illegal harvest is most likely an underestimate,
yet accounts for about 15 percent of total annual catch (Horn et al.
2022; Pauly et al. 2020). Queen conch meat production shows a negative
trend over time and the decrease can largely be attributed to
overfishing (Prada et al. 2017). Some stocks have collapsed and have
yet to recover (Theile 2005; Aldana-Aranda et al. 2003; Appeldoorn
1994b).
Queen conch shells are also used as curios and in jewelry, but are
generally of secondary economic importance. Shells may be offered to
tourists in its natural or polished form (Prada et al. 2017). The large
pinkish queen conch shells are brought to landing sites in only a few
places. In most cases, shells are discarded at sea, generating several
underwater sites with piles of empty conch shells. According to Theile
(2001) from 1992 to 1999, a total of 1,628,436 individual queen conch
shells, plus 131,275 kg of shells were recorded in international trade.
Assuming that each queen conch shell weighs between 700 and 1500 g, the
total reported volume of conch shells from 1992 to 1999 may have been
equivalent to between 1,720,000 and 1,816,000 shells (Prada et al.
2017). In addition, queen conch pearls are valuable and rare, but their
production and trade remain largely unknown across the region. In
Colombia, one of the few jurisdictions with relevant data, exports of
4,074 pearls, valued around USD 2.2 million, were reported between 2000
and 2003 (Prada et al. 2009). With the reduction of the fishing effort
in Colombia, the number of exported queen conch pearls declined from
732 units in 2000 to 123 units in 2010 (Castro-Gonz[aacute]lez et al.
2011). Japan, Switzerland, and the United States are the main queen
conch pearl importers (Prada et al. 2017). Lastly, in recent years,
operculum trade has developed, but similarly little is known about it.
China is the major importer and it is believed opercula are used in
traditional Chinese medicine. In 2020, the U.S. Fish and Wildlife
Service (USFWS) confiscated a shipment in-transit from Miami, Florida
to China (weighing 1 mt) of conch products, consisting largely of
opercula. The shipment was confiscated by USFWS for CITES and U.S.
Lacey Act violations (GCFINET, June 10, 2020).
Indications of Overutilization
In broad terms, a sustainable fishery is based on fishing ``excess
production'' and supported by a stable standing stock or population. In
a sustainable fishery, the abundance of the fished population is not
diminished by fishing (i.e., new production replaces the portion of the
population removed by fishing). Under ideal conditions, the age
structure of a fished population is also stable, for example, without
truncation of the largest, most productive members of the
[[Page 55210]]
population. There are a variety of indications when a fishery resources
is overutilized. Declines in fishing catches or landings with the same
amount of fishing effort (i.e., CPUE) can indicate a population is
being over-utilized. Similarly, changes in spatial distribution (e.g.,
depletions near fishing centers or depletions in more easily accessible
shallow water habitats) likely indicate overutilization. Additionally,
a reduction of genetic diversity or a reduction in maximum size
achieved can indicate severe overutilization. Drastic differences
between population densities found in protected, non-fishing reserves
and those found in fishing areas can also indicate overutilization,
even though the reserve may serve to moderate the effects of
overutilization to a certain extent. These factors were all considered
in the SRT's assessment of the threat and impact of overutilization on
the status of the queen conch. Reductions in distribution as well as
overall population levels can be especially problematic for queen conch
because they require a minimum local adult density to support
reproductive activity.
In particular, available density estimates provide an initial
indication that queen conch may be suffering from overutilization.
Approximately 25 (of 39) jurisdictions have adult conch densities below
the minimum cross-shelf density (50 adult conch/ha) at which
reproductive activity largely ceases. It should be noted, however, that
this minimum density pertains to density within reproductive
populations and not necessarily cross-shelf densities. Overall,
however, the available data suggest that queen conch has been
significantly depleted throughout its range with only a few exceptions.
The jurisdictions of Saba, St. Lucia, Colombia's Serrana Bank,
Nicaragua, Jamaica's Pedro Bank, Costa Rica, Cuba, The Bahamas' Cay Sal
Bank and Jumentos and Ragged Cays, and Turks and Caicos are the only
jurisdictions that have cross-shelf densities above the 100 adult
conch/ha threshold to support reproductive activity resulting in
population growth discussed above. It is likely that populations
residing in inaccessible areas (difficult to fish) may support some
level of mating success and therefore recruitment. However, in these
jurisdictions surveys are not comprehensively performed, and there is
evidence of local overutilization of some populations.
The Landings Data
The SRT evaluated landings data from two international databases.
The FAO maintains data supplied by member nations in their FishStat
database. The queen conch data represent the landings of commercial
fisheries, generally artisanal and industrial, in the Western Tropical
Atlantic; however, discussions are continuing among scientific working
groups regarding the inadequacy and inconsistency of reporting in this
database (FAO Western Central Atlantic Fishery Commission 2020). For
example, the reports from each jurisdiction vary depending on how much
processing has been done (FAO Western Central Atlantic Fishery
Commission 2020). Data are reported either in live weight, which
equates to whole animals, or in various grades of cleaned weight (e.g.,
dirty conch (unprocessed, removed from shell), 50 percent (operculum
and viscera removed), 65 percent (operculum, viscera, and ``head''
(i.e., eyes, stalks, and proboscis) removed), 85 percent (all of the
above plus verge, mantle, and part of the skin removed) and 100 percent
cleaned (fillet, i.e., only the pure white meat remains)). The types of
submitted landings have not always been clearly defined and there is a
continuing effort to encourage jurisdictions to submit consistent queen
conch fisheries data and use standardized conversion factors so data
from different reports can be compared more reliably (FAO Western
Central Atlantic Fishery Commission 2020).
Additional complications in interpreting FishStat data relate to
unexplained changes in local conditions or influences on the fisheries.
Interannual changes in landings may be due to changes in availability
of queen conch (i.e., lowered CPUE), but they may also be due to
changes in regulations or enforcement or unfavorable environmental
conditions (e.g., hurricane disruptions of fishing). Without some
concomitant data on fishing effort, it is difficult to interpret
changing landings.
The second international repository of conch data is maintained by
the CITES. The CITES database records exports and imports of
internationally traded queen conch. The CITES data do not include
commercial catches for local markets and can suffer from many of the
same shortcomings as the FAO FishStat data. Neither database includes
spatial information that allows analysis of local effects on
populations. In addition to providing data for international
obligations, most jurisdictions have widely varying capabilities for
collecting complete data that would adequately characterize all fishing
sectors. They primarily have focused on commercial fishing, either
industrial or artisanal. Jurisdictions have typically inadequately
recorded data from the artisanal commercial fishing sector since
landing sites can be too numerous to effectively monitor with the
limited number of fishing inspectors employed, and self-reporting is
often incomplete. Generally, information is lacking from most
jurisdiction throughout the Caribbean region on recreational or
subsistence fishing, which includes sectors that generally fish for
personal consumption, as well as minor sales or barter of catches. Gaps
also occur in some data collected on catches destined for local
consumption, either by family, neighbors or restaurants. An additional
complication with interpreting ecological and fishery independent data
is that different metrics tend to be used. Commercial landings are
reported in weight and ecological surveys typically count numbers and
estimate or measure lengths of queen conch. Conversion factors may be
jurisdiction- or site-specific, so comparing reported landings to
density surveys has inherent difficulties and opportunities for
miscalculation.
In an effort to fill the gaps in total reported queen conch
landings, the SAU program (Fisheries Centre, Univ. of British Columbia,
<a href="http://www.seaaroundus.org">www.seaaroundus.org</a>) developed a protocol to reconstruct landings
histories for most of the jurisdictions where queen conch is fished.
The SAU scientists assembled available data on landings, supplemented
with additional sociological and fishing data and identified
alternative information sources for missing data by consulting with
local experts and additional literature, to produce their best
estimates of total landings from all fishing sectors. The SAU data
includes subsistence fishing, recreational fishing, and small-scale
artisanal fishing that are generally poorly documented by other
sources. For these reasons, the SRT concluded the SAU data are the most
comprehensive and is the best available data for understanding the
magnitude and impact of all fishing pressure including subsistence,
recreational, and artisanal fishing on local stocks of queen conch. The
SRT compared the reconstructed landings from the SAU project (Pauly et
al. 2020) to the reported FAO landings for queen conch in the western
Caribbean to examine the magnitude of potential differences (see Figure
14 in Horn et al. 2022). Based on this comparison, early reports of FAO
[[Page 55211]]
landings were greatly underestimated. From 1950-59, unreported landings
averaged 93.8 percent of the total SAU-reconstructed queen conch
landings (see Figure 14 in Horn et al. 2022). For regional landings,
the mean percent of unreported landings varied in each decade, 1960-69:
72.1 percent, 1970-79: 53.0 percent, 1980-89: 42.0 percent, 1990-99:
15.8 percent, 2000-09: 23.0 percent, 2010-16: 23.7 percent. Since about
1990, there were improvements in the correlation between FAO and the
SAU-reconstructed landings (ranging from 15-25 percent unreported), but
the FAO landings are unlikely to include all of the fishing sectors in
each jurisdiction, for the reasons discussed above.
To provide a more meaningful comparison with population estimates,
the SAU-reconstructed landings were converted to estimated abundance.
For this region-wide comparison, a standard regional conversion factor
was used (live weight: 1.283 kg/individual, Thiele 2001); subsequent
analyses for specific jurisdictions used location-specific conversion
factors where available. When no jurisdiction or site-specific
information was available, the SRT used the same standard regional
conversion factor. At the peak, regional landings translated into about
32-33 million queen conch per year and, after a slight dip in 2005-
2006, landings remained about 30-31 million queen conch per year from
2012-2016, which is the most recent years with complete data (see
Figure 14 in Horn et al. 2022). Repeatedly in the reports of SAU-
reconstructed landings, the landings are stated as conservative,
underestimating the likely actual landings. The information cited by
the SRT (see S1 in Horn et al. 2022) also provides evidence that many
jurisdictions are landing significant amounts of juvenile or sub-adult
conch, which would be expected to weigh less than 1.283 kg/individual,
thus, the converted abundance figures should also be considered an
underestimation.
The SRT chose to use the SAU-reconstructed landings, when
available, as the best estimate of total landings and used them to
compare exploitation rates (e.g., individuals removed) and stock size
estimates. If SAU-reconstructed landings data were not available, the
SRT used FAO landings data for the comparisons. These data give some
indication of the full magnitude of fishing on queen conch across the
species' range. The mean landings per year from 1950-2016 show that the
12 highest producing jurisdictions have produced 95 percent of the
landings across the region (i.e., Turks and Caicos, The Bahamas,
Honduras, and Jamaica, followed by Belize and Nicaragua, and then
Dominican Republic, Mexico, Cuba, Antigua and Barbuda, Colombia, and
Guadeloupe).
Estimates of Exploitation Rate
Traditional fishery stock assessments use fishery landings data and
indices of relative stock abundance to determine exploitation rates.
However, few jurisdictions collect adequate information (e.g., catch-
per-unit effort data, landings data encompassing all removals) from
their queen conch fisheries to develop traditional stock assessment
models and associated recommendations for sustainable harvest. An
alternative metric using a combination of landings and density surveys
has been recommended by expert working groups and fisheries managers to
estimate exploitation rates. Using this alternative metric, the working
groups and fisheries managers recommend limiting fishing to no more
than 8 percent of mean or median fishable biomass (i.e., standing
stock) as a precautionary sustainable yield, if the stock density can
support successful reproduction (i.e., 100 adult conch/ha) (FAO Western
Central Atlantic Fishery Commission 2013). The 8 percent exploitation
target seeks to ensure that the population per capita growth rate
exceeds the exploitation rate, which in turn ensures population
sustainability under controlled harvest. Using exploitation rates as a
proxy for sustainable yield targets uses fishery-independent estimates
of abundance and fishery-dependent landings data as a substitute for
full stock assessments in data-poor fisheries. Additionally, using
exploitation rates as a proxy depends on statistically valid sampling
to ensure that population extrapolations are an accurate indicator of
population status. This approach also depends on quantifying or mapping
depths and habitats on which to base extrapolations. The FAO also
recommends that the 8 percent exploitation rate be adjusted downward if
the mean conch density is below the level required to support
successful reproductive activity (100 adult conch/ha) (FAO Western
Central Atlantic Fishery Commission 2013).
In an effort to better understand whether adult conch densities can
support current exploitation rates, the SRT plotted the estimated adult
conch densities against recent landings (maximum of either FAO or SAU)
to evaluate regional trends in resource usage (see Figures 18, 19 in
Horn et al. 2022). Exploitation rates for each jurisdiction were
calculated by the SRT as the average numbers landed per year divided by
the total abundance (adults only) across the shelf for the period 2010-
2018 (For additional information on methods, see Horn et al. 2022). The
SRT's analysis suggests that the highest producers in the region,
Dominican Republic, Antigua and Barbuda, Belize, Turks and Caicos, and
Mexico, significantly exceed the 8 percent exploitation rate target.
Additionally, of these jurisdictions, all but Turks and Caicos, have
adult conch densities below the absolute minimum adult density (i.e.,
50 adult conch/ha) required to support any level of reproductive
activity. The fact that these jurisdictions have exceeded the 8 percent
exploitation rate, have adult conch densities below 50 adult conch/ha,
and have not lowered the exploitation rate, indicates harvest is
unsustainable and overutilization is likely occurring. Nicaragua,
Honduras, and Jamaica are fishing near the 8 percent exploitation rate
target. However, while Honduras fishes near the 8 percent exploitation
rate, the adult conch densities are also below the minimum density
threshold (50 adult conch/ha), which also indicates that harvest is
unsustainable and overutilization is likely occurring. The majority of
other conch meat producers within the Caribbean region (e.g., St.
Vincent and the Grenadines, Puerto Rico, Panama, Guadeloupe, Anguilla,
St. Lucia, St. Kitts and Nevis, St. Barthelemy, St. Martin,
Cura[ccedil]ao, U.S. Virgin Islands, and Haiti), are fishing well above
the 8 percent rate and their adult conch densities are well below the
minimum density threshold (50 adult conch/ha), indicating
overutilization is likely occurring. Notably, Aruba, Barbados,
Colombia, The Bahamas, Bonaire, British Virgin Islands, Martinique,
Venezuela, and Grenada, all fish below the 8 percent exploitation rate,
but also have very low adult densities (<50 adult conch/ha), which
suggests that these populations are experiencing recruitment failure
due to depensatory processes, despite the low exploitation rate.
Summary of Findings
Queen conch has been fished in the western tropical Atlantic for
hundreds of years, but in the last four decades, fishing has increased
and industrial scale fishing has developed (CITES 2003). In most
jurisdictions, conch fishing continues although population densities
are very low, with conch populations either experiencing reduced
reproductive activity or having densities so low that reproductive
activity has ceased.
[[Page 55212]]
Several indicators suggest that overfishing is affecting
abundances, densities, spatial distributions, and reproductive outputs
(FAO 2007). In addition, many jurisdictions cite the loss of queen
conch from shallow waters and the need for their fisheries to pursue
conch with SCUBA or hookah in deeper waters (see S1 in Horn et al.
2022).
Efforts to assess the status of queen conch across its range are
hampered by the lack of data collection for all fishing sectors. While
many jurisdictions make an effort to collect data on the main
commercial fisheries, including both industrial and artisanal, the
collections are difficult in artisanal conch fisheries. Artisanal
fisheries typically land queen conch at a wide variety of locations,
lack adequate centralized marketing outlets that can be monitored as a
check on landings, and lack enforcement resources to ensure compliance
with size, quotas, and other regulations. To cope with the short-
comings of data collection, the SAU project implemented an approach to
reconstruct catches for most of the jurisdictions where queen conch is
fished. The SRT relied on these reconstructed landings as best
available scientific information to examine changes in landings over
time and comparisons of landings with standing stock.
The results from the SRT's analysis provide substantial evidence
indicating that overutilization is occurring throughout the species'
range. Only 10 percent (4 jurisdictions) of the 39 jurisdictions
reviewed are fishing at or below the 8 percent exploitation rate and
have adult conch densities that are capable of supporting successful
reproduction (>100 conch/ha), and therefore recruitment (Horn et al.
2022). Forty-one percent of the jurisdictions reviewed are exceeding
the 8 percent exploitation rate and have a median conch densities below
the 100 adult conch/ha threshold required for successful reproductive
activity, while 33 percent of the jurisdictions reviewed are exceeding
the 8 percent exploitation rate and have median conch densities below
the minimum threshold required to support any reproductive activity
(<50 adult conch/ha). Thus, the best available commercial and
scientific information indicates that exploitation levels have resulted
in the overutilization of the species throughout its range and
represents the most significant threat to species.
Inadequacy of Existing Regulatory Mechanisms
The SRT evaluated each jurisdiction's regulations specific to queen
conch, including fisheries management, implementation and enforcement,
to determine the adequacy of existing regulatory mechanisms in
controlling the main threat of overutilization of the species
throughout its range. The SRT identified some common minimum size
regulations that are intended to restrict legal harvest with some form
of size-related criterion. The general goal of the size restrictions is
to offer protection to at least some proportion of queen conch (e.g.,
juveniles or immature conch) that are not yet sexually mature to
preserve reproductive potential. A more detailed summary that includes
the best available information on queen conch populations, fisheries,
and their management in each jurisdictions is presented in its entirety
in the status review report (see S1 in Horn et al. 2022).
Common Queen Conch Minimum Size Regulations
Minimum size regulations are often implemented to help prevent the
harvest of juvenile or immature conch. These minimum size requirements
rely on lip thickness, lip flare, shell length, and meat weight as
indicators of maturity.
Lip thickness is the most reliable indicator for maturity in queen
conch. The best available information indicates that shell lip
thickness for mature queen conch ranges from 17.5 to 26.2 mm for
females, and 13 to 24 mm for males (Stoner et al. 2012; Bissada 2011;
Aldana-Aranda and Frenkiel 2007; Avila-Poveda and Barqueiro-Cardenas
2006). Boman et al. (2018) suggested that a 15 mm minimum lip thickness
would be appropriate for most of the Caribbean region. The primary goal
of a minimum lip thickness is that queen conch will have at least one
season after reaching sexual maturity to mate and spawn. However, many
of the lip thickness requirements discussed below are set too low to
ensure the maturity of the harvested conch.
Regulations that simply require a flared lip to be harvested are
based on a long-outdated idea that maturity occurs at the time of the
flared lip develops (Stoner et al. 2021). Flared shell lips are an
unreliable independent indicator of maturity because as discussed
above, the shell lip can flare a full reproductive season before an
individual can mate or spawn. Similarly, it is well established that
shell length is a poor predictor of maturity in queen conch because
maturity occurs following the termination of growth in shell length,
and final shell length is highly variable with location and
environmental conditions (Tewfik et al. 2019; Appeldoorn et al. 2017;
Foley and Takahashi 2017; Stoner et al. 2012c; Buckland 1989 Appeldoorn
1988a).
Moreover, regulations that impose shell requirements (e.g., shell
length, flared lip or lip thickness) are not enforceable if the shell
is discarded at sea and the conch can be landed out of its shell. Meat
weight is the only maturity measure not associated with the shell and
it is also not a reliable criterion of maturity in queen conch. As
previously discussed, large immature conch can have larger shells
(sometimes with a flared lip) and weigh more than adults. Further, meat
weight requirements that are enforced after the animal is removed from
its shell have reduced effectiveness in limiting the harvest or
protecting reproductive potential because the animal cannot be
returned.
Bermuda
Queen conch were relatively abundant in Bermuda up until the late
1960s, but by the late 1970s populations had reached very low levels
(Sarkis and Ward 2009). Bermuda subsequently closed the queen conch
fishery in 1978 and queen conch is currently listed as endangered under
the Bermuda Protected Species Act 2003. The Bermuda Department of
Conservation Services has developed a recovery plan for queen conch
with the primary goal to promote and enhance self-sustainability of the
queen conch in Bermuda waters. Despite closure of the fishery over 40
years ago, adult densities across the shelf remain low (and below the
50 adult conch/ha required to support any reproductive activity)
suggesting additional regulations or management measures, such as those
aimed at protecting local habitat or water quality, may be warranted.
The SRT's connectivity model (Vaz et al. 2022) indicates that the queen
conch population in Bermuda relies entirely on self-recruitment. Thus,
without management or regulatory measures that not only protect, but
also help grow the adult breeding population, queen conch densities
will likely decline in the future.
Cayman Islands
Concerns about overfishing of queen conch in the Cayman Islands
began in the early 1980s, and in 1988 the Department of Environment
began conducting surveys to monitor the status of queen conch.
Available survey data indicate persistently low queen conch densities
from 1999 to 2006; followed by a decline in 2007 and a modest increase
in 2008 (Bothwell
[[Page 55213]]
2009). The Cayman Islands import the majority of their conch meat, but
there is a small fishery that harvests queen conch for domestic
consumption (Bothwell 2009). The Cayman Islands' 1978 Marine
Conservation Law established a closed fishing season (May 1 through
October 31), during which no conch may be taken from Cayman waters, and
a 5 conch per person or 10 conch per vessel per day bag limit during
the open season. Queen conch fishing is prohibited in Marine Park
Replenishment Zones. There are no minimum size regulations to prevent
harvest of juvenile conch. The use of Self-Contained Underwater
Breathing Apparatus (SCUBA) and hookah diving gear to harvest marine
life is prohibited in the Cayman Islands (Bothwell 2009; Ehrhardt and
Valle-Esquivel 2008). Local Illegal, Unreported, Underreported (IUU)
fishing is a significant issue and regularly occurs in protected areas
by neighboring countries (Bothwell 2009). Given the Caymans' small
shelf area, Bothwell (2009) concluded that even a single poacher, who
requires only simple fishing gear (i.e., mask and fins), can cause
severe problems. In addition to local illegal fishing, the Cayman
Islands also receive IUU queen conch meat fished or exported from
neighboring jurisdictions, and border control has been identified as a
severe weakness (Bothwell 2009). The SRT's connectivity model indicates
(Vaz et al. 2022) that the Cayman Islands are largely a source for
queen conch larvae to other jurisdictions (particularly Cuba), so as
queen conch in the Cayman islands are depleted, other jurisdictions are
less likely to receive recruits from the Cayman Islands (see Figure 12
in Horn et al. 2022). Given the persistently low queen conch densities
over the last decade, lack of minimum size regulations to prevent
juvenile harvest, lack of enforcement, and evidence of significant IUU
fishing, existing regulatory measures within the Cayman Islands are
likely inadequate to protect queen conch from overutilization and
further decline in the future.
Colombia
The queen conch commercial fishery in Colombia shifted to the
continental shelf Archipelago of San Andr[eacute]s, Providencia, and
Santa Catalina (ASPC), including its associated banks
(Quitasue[ntilde]o, Serrana, Serranilla, and Roncador) in the 1970s
when conch populations in San Bernardo and Rosario became severely
depleted due to inadequate regulatory mechanisms (Mora 1994). Even with
the declaration of San Bernardo and Rosario as national parks that
allow subsistence fishing only, densities further declined to very low
levels by 2005 (0.9-12.8 adult conch/ha, 0.2-12.9 juvenile conch/ha),
suggesting recruitment failure (Prada et al. 2009). Prada et al. (2009)
noted that illegal queen conch harvest might represent 2-14 percent of
total harvest (approximately 1.4-21.8 mt of clean meat). During the
1980s and 1990s, a suite of regulatory measures was put in place to
protect populations in the ASPC because it constituted almost all of
Colombia's production. Regulations include area closures, prohibition
on the use of SCUBA gear, a minimum of 225 g meat weight, and a minimum
of 5 mm shell lip thickness (Prada et al. 2009). In addition, the CITES
listing in 1992 established international trade rules. Despite these
measures, fishery-dependent data collected through the mid-1990s and
early 2000s masked continued population declines due to biases
associated with reporting CPUE, incomplete data reporting (e.g.,
inconsistent reporting of landings in versus out of the shell and
incomplete or absent key spatial information), and illegal trade both
into and out of Colombia. For example, in 2008, illegal queen conch
meat exports were traced back to Colombia (as well as other
jurisdictions previously mentioned) during the Operation Shell Game
investigation (U.S. House, Committee on Natural Resources, 2008).
Ultimately, management measures were ineffective as evidenced by
decreased landings, increased effort, and low densities reported by
diver-based visual surveys at two of the three offshore banks: 2.4
conch/ha at Quitasue[ntilde]o and 33.7 conch/ha at Roncador (Valderrama
and Hern[aacute]ndez, 2000). The Colombian government responded by
closing the fisheries at Serrana and Roncador, and reducing the export
quota by 50 percent (CITES 2003). Still these measures were inadequate
and the entire ASPC closed from 2004-2007 due to illegal trade,
conflicts between industrial and artisanal fishers, and discrepancies
between landings and exports (Castro-Gonz[aacute]lez et al. 2009). In
2008 the fishery at ASPC partially reopened at Roncador and Serrana
Banks, with annual production set at 100 mt (Castro-Gonz[aacute]lez et
al. 2011), only to close the fishery at Serrana Bank again in 2012.
The overall adult queen conch densities remain below the critical
threshold required to support any reproductive activity throughout much
of the jurisdiction. Despite very low adult densities (fewer than 50
adult conch/ha in all locations, except at Serrana bank), the queen
conch fishery continues to operate in Colombia. Because the ASPC is
unlikely to receive significant larval input from source populations
outside the area (Vaz et al. 2022), the region may not recover with
current regulatory measures without sufficient adult densities in local
populations. The lack of information for populations in deeper areas
throughout the ASPC, which may be particularly important for recovery
(Castro et al. 2011 unpublished), hinders Colombia's ability to make
comprehensive management decisions and illegal fishing continues to
plague the region. Furthermore, while regulations require a minimum
shell lip thickness of 5 mm and shell lip thickness is a reliable
indicator for maturity in queen conch, this value is likely too low to
protect immature queen conch harvest. Finally, when the shell is
discarded at sea the lip thickness requirement is not enforceable, and
any protective value of the meat weight regulations is diminished.
Costa Rica
Queen conch harvest in Costa Rica was prohibited in 1989 (CITES
2003; Mora 2012). In 2000, the commercial sale of incidentally captured
queen conch was also prohibited, but queen conch caught as bycatch
could be kept for personal consumption. Population declines were
reported in 2001, but there is limited information available related to
those declines (CITES 2003). The adequacy of existing regulatory
measures in protecting queen conch from threats, such as IUU fishing is
unknown.
Cuba
The current status of queen conch populations in Cuba is
questionable due to a lack of available information; however, the few
published surveys suggest relatively high densities, particularly in
protected national parks (e.g., Jardines de la Reina National Park:
1,108 conch/ha in 2005; Formoso et al. 2007; National Park Desembarco
del Granma: 511 conch/ha to 1,723 conch/ha in 2009 to 2010; Cala et al.
2013). The SRT was unable to locate more recent population assessments
or surveys. The commercial harvest of queen conch began in Cuba in the
1960s and the harvest level increased considerably in the mid to late
1970s. However, due to the largely unregulated and unmanaged harvest,
the queen conch population collapsed, and the fishery was closed in
1978. It reopened in the 1982 with a 555 mt harvest quota, which
increased to 780 mt in 1984 (Munoz et al. 1987). Conch populations
continued to decrease at an accelerated
[[Page 55214]]
rate despite the newly established quota system and size based
regulations (Grau and Alcolado as cited in Munoz et al. 1987). Munoz et
al. (1987) attributed the continued population declines to harvest
quotas being set too high and illegal harvest. In 1998 the fishery was
closed again for a year to conduct an abundance survey (Formoso 2001)
and update quotas. Since then, the queen conch fishery has been managed
under a catch quota system that is established by ``zones'' and set
between 15 and 20 percent of the adult queen conch biomass, according
to population assessments and monitoring. The most recent FAO landings
data indicates that queen conch landings have ranged from 475 mt landed
in 2018, 405 mt in 2017, and 477 mt in 2016 (see S2 in Horn et al.
2022); however, no population assessments or surveys were available for
these years. The regulations also include seasonal closures that co-
occur with peak spawning, depth limits on diving operations, a
prohibition on SCUBA gear, and a minimum lip thickness of greater than
10 mm. While shell lip thickness is a reliable indicator for maturity
in queen conch, the minimum 10 mm shell lip thickness regulation likely
does not prevent the harvest of immature queen conch. Additionally,
compliance and enforcement of these regulations appears to be a
problem. For example, two fishing ``zones'' were closed in 2012 because
fishermen were not complying with the regulatory requirements (FAO
Western Central Atlantic Fishery Commission 2013).
Despite the lack of available information on illegal harvest of
conch in Cuba, there is evidence that some limited illegal conch
harvest likely occurs. A recent news article estimated that around one
thousand vessels involving approximately 2,500 people were engaged in
the illegal harvest of marine species, including conch, lobster, and
shrimp (14ymedio 2019). In 2019, Cuba passed new fishery laws aimed at
curbing illegal fishing by instituting a new licensing system (14ymedio
2019). There is currently no information available on the
implementation and enforcement of these new regulations, and the only
survey data available are from surveys of protected areas in 2009. In
addition, Cuba's regulations are meant to implement a catch quota
system that is based on adult biomass estimates, which are obtained
through population assessment, and the most recent population
assessments available are more than 10 years old. Without additional
information on the status of the queen conch population in Cuba or the
effectiveness of the new regulations, the adequacy of existing
regulations is unknown. However, given the history of the conch
fishery, including the rate at which declines can occur with
unsustainable quotas, and the rate of illegal harvest, effective
enforcement of existing regulations, particularly in the protected
areas, is important to protect the queen conch in Cuba from
overutilization in the future.
Dominican Republic and Haiti
Queen conch in the Dominican Republic and Haiti have been
overfished since the 1970s (Wood 2010; Mateo P[eacute]rez and Tejeda
2008; Brownell and Stevely 1981). In 2003, Haiti established
regulations that include a ban on harvesting queen conch without a
flared lip, and the use of SCUBA and hookah gears (CITES 2003).
However, the available information indicates that queen conch are still
fished in Haiti using SCUBA gear (FAO 2020; Wood 2010). Similarly,
while the regulations for a closed season from April 1 through
September 30 exist, the available information indicates that
enforcement is limited (FAO 2020).
The Dominican Republic established regulations for a minimum shell
size in 1986, a closed season in 1999, and no fishing areas in 2002.
But these regulations are reported to be ineffective due to inadequate
enforcement (CITES 2003, 2012). Illegal trade is also common. For
example, from 1999 to 2001, the Dominican Republic almost doubled its
queen conch production, elevating concerns about illegal fishing, which
resulted in the imposition of a CITES moratorium. More recently, in
2008, both Haiti and the Dominican Republic, in addition to Jamaica,
Honduras, and Colombia, were implicated in illegal exports of more than
119 mt of queen conch meat during the Operation Shell Game
investigation (Congress, U.S. House, Committee on Natural Resources,
2008).
Although dated (i.e., more than 10 years old), the available
information indicates that adult queen conch densities are below the
minimum density threshold for any reproductive activity (50 adult
conch/ha). The status of queen conch in the Dominican Republic is
concerning because under historical conditions it likely functioned as
an important ecological corridor, facilitating species connectivity
throughout the region (Vaz et al. 2022). Although there is evidence
that the rates of decline may have slowed in some areas since 2000
(Torres and Sullivan-Sealey 2002) and that some locations have
reproductive activity (Wood 2010), there is no evidence that
regulations have been effectively implemented or enforced (CITES 2003,
2012; Wood 2010; Figueroa and Gonz[aacute]lez 2012). In addition,
detailed, accurate, consistent, and unbiased reporting of fisheries
data is a challenge and creates a barrier to recognizing and
understanding the current status of populations (FAO Western Central
Atlantic Fishery Commission 2020). Thus, the SRT concluded that adult
queen conch densities are well below what is required for healthy
spawning populations at most locations (Posada et al. 1999; Wood 2010)
and continued declines may be irreversible without human intervention
even if fishing pressure is significantly reduced or halted (Torres and
Sullivan-Sealey 2002). Based on the foregoing, existing regulations are
likely inadequate to address the threat of overutilization and reverse
the decline of populations in the Dominican Republic and Haiti.
Jamaica
Jamaica has been a major producer for the queen conch fishery since
the 1990s (Aiken et al. 1999; Appeldoorn 1994a; Prada et al. 2009). The
commercial fishery is focused around Pedro Bank, located approximately
80 km southwest of Jamaica. Fisheries-independent diver-based surveys
began on Pedro Bank in 1994 and these surveys have helped establish
total allowable catch (TAC) limits for the fishery. Queen conch surveys
are conducted about every 3 to 4 years (e.g., 1994, 1997, 2002, 2007,
2011, 2015, and 2018). Queen conch density estimates for all life
stages and depth strata from 1994 to 2018 have remained at a level that
supports successful reproductive activity (142-203 conch/ha; NEPA
2020). However, surveys in 2018 recorded low enough densities (203
conch/ha, age classes was not provided) such that the National
Fisheries Authority of Jamaica implemented a closure of the queen conch
fishery from 2019 to 2020. Due to the lack of funding to conduct a new
survey, the closure was extended to February 2021 (Jamaica Gleaner, Ban
on Conch Fishing Extended to February 2021, April 6, 2020).
In 1994 the queen conch fishery management plan established
guidelines for management measures including a national TAC and
individual quota system (Morris 2012), a closed commercial season
generally extends from August 1 through February 28 (FAO 2022), and a
prohibition on fishing queen conch at depths greater than 30 m (Morris
2012). These regulations are intended to conserve nursery and breeding
areas as well as
[[Page 55215]]
deep spawning stocks (Morris 2012). There are no minimum size based
regulations to prevent harvest of immature conch. There is no closed
season for the recreational fishery, but harvesting is limited to three
conch per person per day (CITES 2003). Currently, annual quotas for
Pedro Bank are determined through a control rule based on harvesting 8
percent of the estimated exploitable biomass (Smikle 2010). Under this
scenario, the maximum catch is fixed when densities are above 100 adult
conch/ha and are progressively reduced if the population density is
reduced. Quotas cannot be increased unless supported by the results of
an in-water survey; however, quotas can be lowered if there is evidence
of problems, such as a drop in catch per unit effort or a survey
indicating a lack of juveniles for future recruitment, and field
surveys are mandated at regular intervals. Additional management
measures include the designation of the South West Cay Special
Fisheries Conservation Area (SWCSFCA) in 2012. Queen conch fishing is
prohibited within the SWCSFCA, which extends in a 2-km radius around
Bird Key on Pedro Bank. Even so, regulations have not been able to
address illegal fishing, which is thought to be problematic based on a
spike in catch statistics reported by Honduras and the Dominican
Republic during two discrete periods between 2000 and 2002 when
Jamaica's fishery on Pedro Bank was closed (CITES 2012). According to
the FAO Western Central Atlantic Fishery Commission (2020), a Jamaican
national fisheries authority was established, but had an unfunded
compliance branch that receives assistance from the Jamaican Coast
Guard and Marine Police, though fisheries issues are not a priority.
Thus, illegal fishing is thought to remain a serious problem, as
further evidenced by the FAO Western Central Atlantic Fishery
Commission (2020) observation that ``. . . there is intense IUU fishing
by vessels from jurisdictions such as Honduras, Dominican Republic and
Nicaragua'' within the large Jamaican EEZ.
Effective conservation management measures are particularly
important for the Pedro Bank queen conch fishery because it is
geographically isolated and receives little gene flow from external
areas. Thus, the future of Pedro Bank's queen conch fishery likely
depends on local recruitment for sustaining its stocks (Kitson-Walters
et al. 2018). The health of the Pedro Bank conch population may also be
important to species connectivity throughout the Caribbean region, as
Jamaica has been identified as an important ecological corridor and a
source of larvae to down current jurisdictions (Vaz et al. 2022).
In summary, management actions to date have maintained queen conch
populations on Pedro Bank, on average, at levels above the necessary
threshold required to support successful reproduction (i.e., greater
than 100 adult conch/ha); however, existing regulations do not protect
immature conch from harvest and may not be adequate to control illegal
fishing, prevent habitat degradation, or reverse the decline of queen
conch in shallower areas.
Leeward Antilles (Aruba, Cura[ccedil]ao, and Bonaire)
No historical or current fisheries data from the Leeward Antilles
islands are available. However, in Bonaire, Lac Bay historically was
considered to have been ``plentiful in conch.'' (STINPA 2019, as cited
in Patitas 2010). Fisheries were closed in Bonaire and Aruba in 1985
and 1987, respectively, but enforcement of the closure did not begin in
Bonaire until the mid-1990s (van Baren 2013). Limited permits, allowing
take of adult conch over 18 cm shell length or meat weight over 225
grams (g), were issued in Bonaire through the 1990s. But a moratorium
on permit issuance was reported in 2012 due to concern over the
extremely low adult population size at that time (van Baren 2013). The
limited fisheries-independent monitoring suggests that the island-wide
density of conch in Bonaire is very low 21.8 conch/ha. Current
densities are too low to support fisheries, despite being closed for
more than 30 years in two of the three islands (i.e., Aruba and
Bonaire). Queen conch are imported legally from Jamaica and Colombia
and illegally from Venezuela to markets in Cura[ccedil]ao and Bonaire
(FAO 2007).
The most recent study to assess the status of queen conch in
Bonaire was conducted in 2010 in Lac Bay (Patitsas 2010). Within Lac
Bay, overall conch density was recorded to be 11.24 conch/ha. The
majority of conchs in Lac Bay were adults, constituting 85 percent of
the total found (Patitsas 2010). The previous conch density study in
Lac Bay was conducted in 1999, and estimated the overall population to
be around 22 conch/ha with an average age of 2.5 years (Lott 2001, as
cited in Patitsas 2010). Patitsas (2010) concluded the densities in Lac
Bay are below the Allee effect threshold of 50 adult conch/ha (Stoner
and Culp 2000). No surveys have been done to determine the density and
the conditions of the populations in the island of Cura[ccedil]ao
(Sanchez, 2017). The only information of the populations in the island
of Cura[ccedil]ao located by the SRT is presented in a 2017 thesis on
the diet and size of queen conch around the island of Cura[ccedil]ao
(Sanchez 2017). While, Sanchez (2017) did not provide conch density
data, the author concluded that adult queen conch are very rare
surrounding the island, and appear to only occur in restricted places,
like the Sea Aquarium Basins, where illegal fishing and predation is
limited (Sanchez 2017). The average density of queen conch on the west
side of Aruba was 11.3 conch/ha from 2009 to 2011, and the population
was dominated by juveniles, suggesting Aruba populations on the west
side of the island are not large enough for successful reproduction,
though there are isolated areas of higher conch densities (Ho 2011).
There is evidence that illegal fishing continues and is further
contributing to declines (van Baren 2013; Ho 2011; FAO 2011).
Despite fisheries closures in Bonaire and Aruba since the 1980s,
the best available information indicates that there has been limited or
no recovery. The most recent available survey, although dated (i.e.,
more than 10 years old) and discussed above, reported very low conch
densities and suggest further decline in Lac Bay, Bonaire. There is
limited evidence of improvements to management, enforcement, and
conservation planning strategies in Aruba, Cura[ccedil]ao, and Bonaire.
The lack of recovery in the respective conch populations despite the
complete closures of the conch fisheries, indicates that the closures
were likely implemented too late because adult conch densities were too
low to support reproductive activity. In addition, Aruba, Curacao, and
Bonaire appear to have historically relied on larval subsidies of local
origin and from Venezuela, and are mostly isolated from other sources
of larval supply. Therefore, their ability to recover post
overutilization is limited.
Leeward Islands (Anguilla, Antigua and Barbuda, British Virgin Islands,
Guadeloupe and Martinique, Montserrat, Saba, St. Barth[eacute]lemy, St.
Martin, St. Eustatius, St. Kitts and Nevis, U.S. Virgin Islands)
Based on the available data, as described in Horn et al. (2022),
indicates that the majority of the Leeward Islands (i.e., Anguilla,
Antigua and Barbuda, British Virgin Islands, Guadeloupe and Martinique,
Montserrat, St. Barth[eacute]lemy, St. Eustatius, St. Martin, St. Kitts
and Nevis, and U.S. Virgin Islands) have queen conch populations that
are overexploited, with estimated population densities that are below
that
[[Page 55216]]
which is necessary for reproductive success (100 adult conch/ha). The
existing regulatory mechanisms largely appear inadequate, resulting in
overexploitation and illegal fishing, and have likely contributed to
the decline in these populations and reproductive failure. For example,
in Anguilla, surveys conducted in 2015 and 2016 found 26 adult conch/
ha, which is well below the minimum density threshold for any
reproductive activity (50 adult conch/ha) and may not be supporting any
reproductive activity (Izioka 2016). Despite low adult densities,
fishing for queen conch is still allowed. In addition, existing
regulatory mechanisms do not prevent immature queen conch from being
harvested. Currently, the minimum landing size for queen conch in
Anguilla is 18 cm shell length; however, Wynne et al. (2016) found that
up to 94 percent of queen conch harvested at that size were immature.
In Antigua and Barbuda, surveys of populations also show low
densities and low proportions of adult conch, suggesting that fishing
pressure has significantly reduced the adult population to the point
where Allee effects are occurring (Ruttenberg et al. 2018; Tewfik et
al. 2001). For example, Tewfik et al. (2001) conducted 34 visual
surveys (12.84 hectares total) off the southwestern side of Antigua.
These surveys recorded 3.7 adult conch/ha, significantly below the 50
adult conch/ha threshold required to support any reproductive activity.
Overall conch density (adults and juveniles) for Antigua were 17.2
conch/ha, with juveniles making up about 78.4 percent of the entire
population. Reported conch densities in Barbuda are also very low.
Ruttenberg et al. (2018) reports 29 <plus-minus> 12 adult conch/ha and
96 <plus-minus> 30 juvenile conch/ha (mean <plus-minus> SE). In terms
of regulations, both jurisdictions prohibit harvesting of queen conch
without a flared lip, or a shell length less than 180 mm, or animals
whose meat is less than 225 g without the digestive gland. In addition,
Horsford (2019) found over 20 percent of landed conch meat samples were
below the minimum legal meat weight in 2018 and 2019, including conch
harvested within marine reserves. Evidence of the harvest of undersized
and immature queen conch suggests that the existing regulations are
either inadequate or are not enforced, or both. Based on the size
distribution of queen conch in Barbuda, existing regulations do not
necessarily prevent harvesting of immature queen conch. In 2003 the
British Virgin Islands implemented regulations that require an 18 cm
minimum shell length, a flared lip, a meat weight of at least 226 g,
and established a closed season (June 1 through September 30) and
prohibited SCUBA gear. However, enforcement of these regulations is
questionable as the fishery appears to be essentially unmonitored (Gore
and Llewellyn 2005). In addition, as previously discussed shell length
and flared shell lip are not reliable indicators of maturity and likely
do not prevent immature queen conch from harvest. Given that surveys of
queen conch populations in 1993 and 2003 both showed densities of queen
conch on the order of less than 0.07 conch/ha, existing regulatory
mechanisms may not adequately protect queen conch in the British Virgin
Islands from overexploitation (CITES 2003; Ehrhardt and Valle-Esquivel
2008; Gore and Llewellyn 2005).
In Guadeloupe and Martinique, demand is high for local consumption
of queen conch (CITES 2003). In 1986, Martinique passed regulations to
prohibit the harvest of queen conch with a shell length of less than 22
cm, or shells without a flared lip, or animals whose meat weighs less
than 250 g. The majority of landings in Martinique are meat only (FAO
2020), which means that immature queen conch can potentially be
harvested as long as the meat weight is greater than 250 g. In
Martinique, a closed season runs from January 1 through June 30, and
the use of SCUBA gear to harvest conch is prohibited. Studies on the
reproductive cycle of queen conch in Martinique and Guadeloupe have
concluded that the minimum shell length size is not an effective
criterion to base sexual maturity (Frenkiel et al. 2009; Reynal et al.
2009). Thus, the best available information indicates that these
regulatory measures are inadequate to prevent the harvest of immature
queen conch. Given the increasing demand, with the price of queen conch
meat having doubled over the past 25 years (FAO 2020; FAO Western
Central Atlantic Fishery Commission 2013), the existing regulations
will likely continue to contribute to harvesting of immature queen
conch and declines in the queen conch population in the future.
The island of Saba supported large conch fisheries until the mid-
1990s. Intensive and unsustainable harvest during the mid-1980s and
throughout the 1990s led to the declines on Saba Bank. The Saba Bank
was also overfished by several foreign vessels (van Baren 2013). In
1996, fishery legislation prohibited the harvest of queen conch for
commercial purposes, and allowed only Saban individuals to harvest
queen conch for private use and consumption. These regulations limit
Saban individuals to no more than 20 conch per person per year and
require that catch be reported to the manager of the Saba Marine Park
(van Baren 2013). Nonetheless, collection and reporting laws are not
enforced (van Baren 2013). Additional regulations require a 19 cm
minimum shell length or a ``well-developed lip,'' and prohibit SCUBA
and hookah gears (van Baren 2013). No surveys have been conducted to
determine the status of queen conch or if the commercial closure has
been effective in rebuilding queen conch stocks (van Baren 2013).
Anecdotal evidence indicates that queen conch on the Saba Bank are
fished by foreign vessels (FAO Western Central Atlantic Fishery
Commission 2013). The island of St. Eustatius had a small commercial
conch fishery that exported to St. Maarten. In 2010 the fishery was
curtailed because St. Maarten began to require CITES permits for their
imports (van Baren 2013).
In the U.S. Virgin Islands, the U.S. Federal government has
jurisdiction within the U.S. Virgin Island EEZ (i.e., those waters from
3-200 nautical miles (4.8-370 km) from the coast) and the CFMC and NMFS
are responsible for management measures for U.S. Caribbean federal
fisheries. The Government of the U.S. Virgin Islands manages marine
resources from the shore out to the 3 nautical miles. At present, the
U.S. Virgin Islands manages fisheries resources cooperatively with the
CFMC, although not all regulations are consistent across the state-
Federal boundary. Recently, the Secretary of Commerce approved three
new fishery management plans (FMP) for the fishery resources managed by
the CFMC in Federal waters of each of St. Thomas, St. John, and St.
Croix. The St. Thomas and St. John FMP and the St. Croix FMP will
transition fisheries management in the respective EEZ from the
historical U.S. Caribbean-wide approach to an island-based approach;
however, this change does not alter existing regulations for the queen
conch fishery. In the U.S. Caribbean EEZ, no person may fish for or
possess a queen conch in or from the EEZ, except from November 1
through May 31 in the area east of 64[deg]34' W longitude which
includes Lang Bank east of St. Croix, U.S. Virgin Islands (50 CFR
622.491(a)). Fishing for queen conch is allowed in territorial waters
of St. Croix, St. Thomas, and St. John from November 1 through May 31,
or until the queen conch annual quota is reached. The annual quota is
22.7 mt (50,000 lbs) for St. Croix territorial
[[Page 55217]]
waters and 22.7 mt (50,000 lbs) for St. Thomas and St. John territorial
waters (combined). The CFMC established a comparable annual catch limit
(ACL) for harvest of queen conch within the EEZ around St. Croix east
of 64[deg]34' W longitude, which includes Lang Bank. When the ACL is
reached or projected to be reached across territorial and Federal
waters, the Federal queen conch fishery within the EEZ around St. Croix
is closed. From 2012 to 2020, commercial fishermen in St. Croix landed
between 24 and 74 percent of their ACL; therefore, there were no
closures of the queen conch fishery during this time period. In
addition to the harvest quotas, commercial trip limits and recreational
bag limits for queen conch harvest apply in both territorial waters and
Federal waters of the U.S. Virgin Islands. The commercial trip limit in
territorial waters and in the U.S. Caribbean EEZ around St. Croix is
200 queen conch per vessel per day (50 CFR 622.495). The recreational
bag limit from the EEZ around St. Croix is three per person per day or,
if more than four persons are aboard, 12 per vessel per day (50 CFR
622.494). The recreational bag limit in territorial waters is six conch
per person per day, not to exceed 24 conch per vessel per day. In the
EEZ around St. Croix and in U.S. Virgin Islands territorial waters,
regulations require a 22.9 cm minimum shell length or 9.5 mm lip
thickness (50 CFR 622.492). In the EEZ around St. Croix and in U.S.
Virgin Islands territorial waters, queen conch must be landed alive
with meat and shell intact. Finally, Federal regulations at 50 CFR
622.490(a) prohibit the harvest of queen conch in the EEZ around St.
Croix by diving while using a device that provides a continuous air
supply from the surface.
Surveys of queen conch were conducted in the U.S. Virgin Islands in
2008-2010. The median cross shelf adult density estimate for the three
island groups is 44 adult conch/ha, suggesting that densities are too
low to support reproductive activity (Horn et al. 2022). However, queen
conch densities (at all the island groups) were higher in 2008 through
2010 than those observed in the 1980s and 1990s (Boulon 1987;
Friedlander 1997; Friedlander et al. 1994; Gordon 2002; Wood and Olsen
1983). For example, the mean adult queen conch density estimated for
St. Thomas was five times that of adult conch in 2001 (24.2 adult
conch/ha) and four times that in 1996 (32.2 adult conch/ha) and ten
times that in 1990 (11.8 adult conch/ha) (Gordon 2010). In the 2008-
2010 surveys, the population was composed mainly of juveniles (greater
than 50 percent) with the remainder of the population spread evenly
among the older age classes. Similarly, a more recent survey conducted
in Buck Island Reef National Monument (a no-take reserve) estimated
68.5 adult conch/ha and 233.5 juvenile conch/ha (Doerr and Hill, 2018).
This age class structure suggests some successful recruitment in this
area. However, due to the age of the data from the 2008-2010 surveys, a
more recent assessment could better inform stock status. NMFS's 2022
second quarter update to its Report to Congress on the Status of U.S.
Fisheries identifies the queen conch stock in the Caribbean as
overfished, but not currently undergoing overfishing.
Overall, while queen conch regulations exist within the Leeward
Islands to prohibit the harvesting of immature queen conch and manage
fisheries, many of these regulations use inadequate proxy measures for
maturity, are poorly enforced, and lack effective monitoring controls.
For example, minimum shell length, flared lip, and meat weight
regulations are unreliable measures to protect immature conch. While
lip thickness is a more reliable indicator of maturity for queen conch,
values set too low do not ensure that only mature conch are harvested
(Doerr and Hill, 2018; Frenkiel et al. 2009; Reynal et al. 2009;
Horsford 2019). The connectivity models (Vaz et al. 2022) show a
reliance on self-recruitment for the Leeward Islands, with larval
transport mainly away from the islands. Thus, queen conch populations
throughout the Leeward Islands may continue to decline in the future
due to the inadequacy of many of the existing regulatory measures in
protecting the Leeward Island conch populations from overutilization
and limited larval supply from other locations.
Nicaragua
In Nicaragua, the queen conch fishery was not considered a major
fishery until the mid 1990s (CITES 2012). The majority of the queen
conch harvest is caught by fishermen targeting lobster, with the
remainder made by divers during the lobster closed season (Barnutty
Navarro and Salvador Castellon 2013) or incidentally (Escoto
Garc[iacute]a 2004). Landings, quotas, and exports have all increased
significantly since the 1990s (S[aacute]nchez Baquero 2009). In 2003,
Nicaragua implemented regulations that established a 20 cm minimum
shell length, a minimal lip thickness of 9.5 mm, a seasonal closure
from June 1 through September 30, and set the export quota at 45 mt
(Barnutty Navarro and Salvador Castellon 2013; FAO Western Central
Atlantic Fishery Commission 2020). Since then, the export quota has
increased significantly. In 2009, the export quota was set at 341 mt of
clean fillet and 41 mt for research purposes. In 2012, Nicaragua gained
additional conch fishing grounds through the resolution of a maritime
dispute with Honduras (International Court of Justice 2012), and
increased its export quota to 345 mt (Barnutty Navarro and Salvador
Castellon 2013; FAO Western Central Atlantic Fishery Commission 2013).
By 2019, this quota had almost doubled to an annual export quota of 638
mt (FAO Western Central Atlantic Fishery Commission 2020). The 2020
export quota increased again to 680 mt (see CITES Export Quota).
Whether these regulations are adequate to protect the queen conch
population from overexploitation is unclear, but a comparison of queen
conch densities over the years suggests the current quota may be set
too high. For example, results from a 2009 systematic cross-shelf
scientific survey conducted by SCUBA divers showed densities ranging
from 176-267 adult conch/ha depending on the month (April, July, or
November), location, and depth (10-30 m) (Barnutty Navarro and Salvador
Castellon 2013). More recent surveys, conducted in October 2016, March
2018, and October 2019, show a decrease in densities to 70-109 conch/ha
(FAO Western Central Atlantic Fishery Commission 2020). However,
details on these surveys were unavailable and it is unclear if these
are adult queen conch densities. Regardless, the available information
suggests that overall densities have decreased substantially since
2009, presumably due to the significant increases in the export quota
over the past few years. While the densities, if they reflect adult
conch densities, may still support some reproductive activity within
the queen conch population, the existing regulatory measures, including
the current quota, may not be adequate to prevent further queen conch
declines in the future. If these trends continue this population is
vulnerable to collapse, as the connectivity model (Vaz et al. 2022)
indicates that Nicaragua's queen conch population is mostly reliant on
self-recruitment.
Panama
There is little information available on the status of queen conch
or harvest of queen conch in Panama. Georges et al. (2010) suggested
that the queen conch fishery in Panama may not have specific
regulations, but recognized harvest using SCUBA gear is prohibited. In
the 1970s, a subsistence fishery was
[[Page 55218]]
centered in the San Blas Islands (Brownell and Stevely 1981). By the
late 1990s, landings data suggest that the queen conch population had
collapsed (CITES 2003; Georges et al. 2010). In 2000, extremely low
adult densities were observed at Bocas del Toro archipelago
(approximately 0.2 conch/ha; CITES 2003). The most recent information,
although dated, indicates that the fishery was closed for 5 years in
2004 (CITES 2012) and a ``permanent closed season'' remains in place as
of 2019 (FAO 2019). The SAU data suggests that queen conch harvest has
continued during the closure with unreported landings likely occurring
for subsistence and by the artisan fishery (Pauly et al. 2020). In
Panama, queen conch appear to be largely self-recruiting (Vaz et al.
2022) and more vulnerable to depletion as the population likely does
not receive larval recruits from other jurisdictions. The best
available information suggests that Panama does not have adequate
regulatory measures in place to manage queen conch harvest. While it
appears that the harvest is limited to subsistence, the available
information suggests that the population has collapsed, and without
additional regulations and appropriate conservation planning, it is
unlikely that Panama's severely depleted queen conch population will
recover.
Puerto Rico
Queen conch populations in Puerto Rico showed signs of steady
decline beginning in the 1980s (CITES 2012). Estimated fishing
mortality exceeded estimates of natural mortality, catch continued to
decline while effort increased through 2011 (CITES 2012), and the catch
became increasingly skewed to smaller sizes, all suggesting that Puerto
Rican populations have been overfished for decades (Appeldoorn 1993;
SEDAR 2007). Surveys conducted in 2013 observed larger size
distributions, higher adult queen conch densities (compared to three
previous studies, but lower than the density reported in 2006), an
increase in the proportion of older adults, and evidence of sustained
recruitment, suggesting that Puerto Rico's conch populations are
recovering to some extent (Jim[eacute]nez 2007, Baker et al. 2016).
There are several regulations associated with the Queen Conch
Resources Fishery Management Plan of Puerto Rico and the U.S. Virgin
Islands (CFMC 1996). Recently, the Secretary of Commerce approved new
FMPs for the fishery resources managed by the CFMC in Federal waters of
U.S. Caribbean. The Puerto Rico FMP will transition fisheries
management to an island-based approach.
In 1997, the U.S. Caribbean EEZ (with the exception of St. Croix,
U.S. Virgin Islands) was closed to queen conch fishing and a closed
season (July 1 through September 30) for territorial waters was
implemented. In 2004, additional regulations implemented in local
waters included a 22.86 cm minimum shell length or a 9.5 mm minimum lip
thickness requirement, daily bag limits of 150 per person and 450 per
boat, and a requirement to land queen conch intact in the shell. In
2012, the territorial waters seasonal closure was amended to begin on
August 1 and extend until October 31.
In 2013, the Puerto Rico Department of Natural Resources
implemented an administrative order that lifted the prohibition on
extracting conch meat from the shell while underwater (Puerto Rico
Department of Natural and Environmental Resources Administrative Order
2013-14). The administrative order remains valid today. The elimination
of an important accountability mechanism to ensure compliance and
enforcement with the minimum size regulations (i.e., the requirement
that conch be landed whole), occurred while populations were still
considered severely depleted and subjected to continued fishing
pressure. Furthermore, shell length is not a reliable indicator of
maturity in queen conch. As previously discussed, shell lip thickness
is the most reliable indicator of maturity in queen conch; however, the
available information indicates that the 9.5 mm lip thickness
regulation is not high enough to prevent immature conch from being
harvested. Lastly, the mesophotic reef off the west coast of Puerto
Rico is likely an important ecological corridor for maintaining
connectivity between the Windward Islands and the western Caribbean
(Vaz et al. 2022; Truelove et al. 2017), which means that a decline in
queen conch could implicate other jurisdictions down-current. Based on
the foregoing, existing regulations are likely inadequate to reverse
the decline of queen conch in Puerto Rico.
The Bahamas
Landings data from the 1950s through 2018 have ranged between
approximately 750-6,000 mt, with a steadily increasing trend over that
period. Prior to 1992, the export of queen conch from The Bahamas was
illegal. More recently, at least 51 percent of the landings are
exported, with export amounts and values increasing over time, and the
bulk of the product exported (99 percent) going to the United States
(Posada et al. 1997, Gittens and Braynen 2012). The Bahamian government
began implementing an export quota system in 1995 and more recently
additional protective measures have been implemented including: a SCUBA
ban, limited use of compressed air, establishment of a network of
marine protected areas, and restricting take to conch with well-formed
flared lips (FAO 2007; Gittens and Braynen 2012). The Bahamas also
established closed areas, but not closed seasons (Prada et al. 2017).
Concerns continue regarding IUU fishing, which is likely exacerbating
the serial depletion that queen conch are experiencing throughout most
of The Bahamas (Stoner et al. 2019).
Several fishery-independent studies in both fished and unfished
areas within The Bahamas have reported one or more of the following
trends since the late 1990s: declines in adult queen conch densities, a
reduction in the size of adults on mating grounds, a reduction in the
average age of individuals within populations, and a reduction in the
number of immature queen conch within nursery grounds (Stoner et al.
2019). Recent surveys suggest adult queen conch densities are too low
to support any reproductive activity (i.e., <50 adult conch/ha), except
in the most remote areas (Stoner et al. 2019). Substantial decreases in
adult conch densities (up to 74 percent) observed in repeated surveys
in three fishing grounds indicate that the conch population is
collapsing. In fact, Stoner et al. (2019) found that only one location
of the 17 locations surveyed in 2011 and 2018, had reproductively-
viable adult conch densities. Declines in juvenile populations were
reported near Lee Stocking Island where aggregations associated with
nursery grounds were estimated to have decreased by more than half
between surveys conducted in the early 1990s and 2011 (Stoner et al.
2011; Stoner et al. 2019). Visual surveys spanning two decades show
that densities of adult queen conch had a significant negative
relationship with an index of fishing pressure. These surveys also
reveal that average shell length in a population was not related to
fishing pressure, but that shell lip thickness declined significantly
with fishing pressure (Stoner et al. 2019). Other less quantitative
observations on changing queen conch populations, have been observed
over the decades in several nursery grounds (e.g., Vigilant Cay and
Bird Cay). While, juvenile aggregations are subject to large inter-
annual shifts in conch recruitment (Stoner 2003), these
[[Page 55219]]
nurseries are typically inhabited by three year classes or more at any
one time. However, the near total loss of queen conch at these sites
indicates a multi-year recruitment failure or heavy illegal fishing on
the nursery grounds (Stoner et al. 2019; Stoner et al. 2009).
Densities have also declined significantly in three repeated
surveys conducted over 22 years in a large no-take fishery reserve
(Stoner et al. 2019). Unlike fished populations, the protected
population has aged and appears to be declining because of lack of
recruitment (Stoner et al. 2019). Queen conch populations around Andros
Island, the Berry Islands, Cape Eleuthera, and Exuma Cays are at or
below critical densities for successful reproduction (i.e., >100 adult
conch/ha). A fishery closure in the Exuma Cays Land and Sea Park since
1986 has been ineffective in reversing the collapse of the stock in
this area (Stoner et al. 2019). Some areas of the southern Bahamas,
including Cay Sal and Jumentos and Ragged Cays, have maintained queen
conch densities greater than 100 adult conch/ha (Souza Jr. and Kough
2020; Stoner et al. 2019). However, fishing grounds in the central and
northern Bahamas, including the Western and Central Great Bahamian
Banks and Little Bahamian Bank, are depleted and regulatory measures
are needed to reverse the downward trend (Souza and Kough 2020). Media
reports from 2010 through 2020 indicate that remote Bahamian banks are
increasingly threatened by illegal fishing as fishers deplete more
accessible areas (Souza Jr. and Kough 2020).
The Bahamas is largely self-recruiting, retaining the majority of
conch larvae (Vaz et al. 2022). The Bahamas does not export a
significant amount of larvae to most jurisdictions; however, it does
receive a substantial amount of larvae from Turks and Caicos, and to a
lesser extent Cuba (Vaz et al. 2022). The sustainability of queen conch
populations in The Bahamas relies heavily on domestic regulations.
Based on the foregoing, the current status and trends of queen conch in
The Bahamas indicates that existing regulatory measures in The Bahamas
are inadequate to protect queen conch from overutilization and further
declines.
Turks and Caicos
The Turks and Caicos one of the largest producers of queen conch
meat, providing roughly 35 percent of the total landings reported for
the Caribbean region from 1950-2016. In 1994, regulatory measures
prohibited the use of SCUBA gear, established annual quotas, set a
minimum shell length of no less than 18 cm or a minimum meat weight of
no less than 225 g, and stated that all conch landed must have a flared
lip. In 2000, a closed season to exports (July 15 through October 15)
was established, although queen conch can still be harvested for local
consumption during the closed season (DEMA 2012). As previously noted,
shell length, flared lip, and meat weight requirements are not reliable
indicators of maturity. The existing regulations do not include a
minimum lip thickness requirement. It is also notable that queen conch
are not required to be landed whole, but the meat may be removed from
the shell at sea (Ulman et al. 2016), which undermines the
effectiveness of most minimum size-based regulations. In addition,
while a closed season to exports may decrease demand during the
species' reproductive season, it does not fully prohibit the harvest of
spawning adult conch.
Two recent studies suggest that the level of exploitation of conch
populations in Turks and Caicos may be higher than previously thought.
The first study by Ulman et al. (2016) performed catch reconstructions
that identified a significant problem with underreported fishery
landings data from 1950 to 2012. The authors found that the total
reconstructed catch was approximately 2.8 times higher than that
reported by the Turks and Caicos to the FAO, and 86 percent higher than
the export-adjusted national reported baseline. The discrepancies arose
because local consumption was not reported and in fact, the total local
consumption of queen conch accounted for almost the entire total
allowable catch before exported amounts were considered. In response to
this study, the catch quota was lowered in 2013.
The last available queen conch survey was completed in 2001. While
dated, this survey recorded queen conch densities at 250 adult conch/ha
(DEMA 2012). Queen conch harvest is prohibited in the Admiral Cockburn
Land and Sea National Park and in the East Harbor Conch and Lobster
Reserve. Both protected areas are located in South Caicos (CITES 2012).
A study by Schultz and Lockhart (2017) examined the demographics of
conch populations inside and outside the East Harbor Conch and Lobster
Reserve. The authors identified a lack of algal plain habitat, smaller
conch, and lower densities of conch in the reserve. Only one of 118
sites examined inside the reserve contained densities of more than 50
adult conch/ha and none of the sites had densities of more than 100
adult conch/ha. Outside of the reserve, only four of 96 sites had
densities of more than 50 adult conch/ha and only one site had a
density of more than 100 adult conch/ha. Overall, the densities inside
and outside the reserve were similar and had declined by at least an
order of magnitude since 2000. The authors cite a lack of habitat
inside the reserve and continued fishing pressure within the reserve
due to low enforcement presence, as the most likely reasons for an
underperformance of the reserve for queen conch conservation.
The Turks and Caicos likely supplies larvae to The Bahamas, and is
unlikely to receive larvae from overfished populations up current, and
is largely self-recruiting (Vaz et al. 2022). Thus, local reproduction
is critical for sustaining queen conch in Turks and Caicos. The Turks
and Caicos has been one of the largest producers of queen conch meat
for decades; however, recent density trends suggest that existing
regulations may be inadequate to sustain viable populations.
United States (Florida)
Within the continental United States, queen conch only occur in
Florida, where the historical queen conch harvest supported both
commercial and recreational fisheries. Regulatory measures were put in
place in the 1970s, 1980s, and 1990s (Florida Administrative Code,
1971, 1985, 1990) to first limit and then prohibit commercial and
recreational take of queen conch in order to reverse the downward trend
of queen conch populations in Florida (Florida Department of State
2021; Glazer and Berg Jr. 1994). The 1990 regulations also provided a
stricter framework for shell possession. Habitat loss resulting from
coastal developmental contributed to the decline of queen conch
populations during the 1980s, and since that time, multiple state and
Federal regulations (e.g., Florida Department of Environmental Planning
and the Florida Keys National Marine Sanctuary) have limited discharge,
development, and other anthropogenic activities that may influence
water quality and degrade coastal habitat.
Queen conch are grouped into three ``subpopulations'' within the
Florida Keys based on their spatial distribution (i.e., nearshore,
back-reef, and deep-water) (Glazer and Delgado 2020). To date, none of
the above measures have been effective in restoring subpopulations in
the nearshore, shallow water, and hard bottom habitats immediately
adjacent to the Florida Keys island chain. In fact, three populations
known to exist in the 1990s remain locally extinct despite 35 years of
fishery closure (Glazer and Delgado 2020). Most queen conch in the
[[Page 55220]]
nearshore areas are not capable of reproduction, which in part, may be
due to deficiencies in their gonadal development (Glazer et al. 2008;
Spade et al. 2010; Delgado et al. 2019), and very low densities. While
the reason for reproductive failure in the nearshore areas has not been
clearly identified, contaminants may also play a role in the
reproductive failure. In addition, low adult densities, high water
temperatures, and natural geographic barriers to movement (e.g., Hawks
Channel) appear to limit opportunities for the formation of spawning
aggregations that could restore viable populations in nearshore areas.
Therefore, it is likely that these populations will continue to decline
without additional intervention, despite the protective measures that
have been in place for 50 years.
The Florida Keys' back-reef subpopulation is located in shallow
water reef flats in habitats primarily consisting of coral rubble,
sand, and seagrass (Glazer and Kidney 2004), and has been the focus of
fishery-independent surveys since 1993 (Delgado and Glazer 2020). These
surveys confirm that the adult abundance of queen conch on back reefs
in the Florida Keys has been increasing slowly but steadily since 2007.
By 2013, with a few setbacks due to major hurricanes in 2004 and 2005,
adult abundance reached approximately 65,000 individuals (Glazer and
Delgado 2020). Delgado and Glazer (2020) have confirmed that adult
spawning densities in the back-reef are high enough (exceeding 100
adult conch/ha) to support successful reproduction, although the
authors never observed mating when aggregation density was less than
204 adult conch/ha, and spawning was not observed when densities were
less 90 adult conch/ha.
In summary, queen conch in Florida have experienced large declines
since the 1970s due to fisheries harvest and habitat degradation,
despite protective regulations being put in place in the 1970s, 1980s,
and 1990s. The best available data indicate that the density of large
adults is still too low and compromised (i.e., non-reproductive adults
in nearshore areas) to restore healthy subpopulations in the Florida
Keys: nearshore, back reef, and deep-water. The median adult queen
conch density in Florida is less than 50 conch/ha, which is too low for
successful reproduction to be maintained throughout the region and for
Florida to have a healthy self-recruiting population. Evidence of
increasing abundance on back reefs and the restoration of the
reproductive capacity of nearshore adult conch following translocation
is promising. Fishery closures and other regulatory measures
implemented up until the early 2000s may be partially responsible for
some of the positive trends that have been observed within the last
decade. Recent restoration measures through translocation implemented
by the State suggest that queen conch populations may have the capacity
to recover with sustained human intervention. Additional regulatory
measures outside of Florida are unlikely to have a positive impact on
queen conch occurring within Florida because connectivity modeling (Vaz
et al. 2022) and genetic analysis (Truelove et al. 2017) suggest that
Florida is largely a self-recruiting population. The commercial and
recreational fishery closures in Florida are likely adequate to prevent
further overutilization, but, given the longevity of the closures and
lack of recovery observed, particularly in nearshore, additional
restoration measures are likely needed.
Venezuela
The commercial conch fishery in Venezuela occurred almost
exclusively in the insular region, with the archipelagos of La Orchila,
Los Roques, Los Testigos, and Las Aves all having significant conch
densities (Schweizer and Posada 2006). Until the mid 1980s queen conch
were predominantly harvested in Los Roques Archipelago. Studies of the
queen conch population around Los Roques Archipelago in the 1980s
(Guevara et al. 1985) showed the population to be severely overfished,
and subsequently the Los Roques Archipelago conch fishery was closed in
1985. Despite the closure, high landings continued (e.g., 360 mt in
1988) and in 1991, the entire commercial queen conch fishery closed
(CITES 2003). Most recently, the FAO reported the following annual
landings data at 2 mt, in 2016, 2017, and 2018 (see S2 in Horn et al.
2022). This illegal harvest of queen conch despite the closure, as well
as illegal fishing by other jurisdictions, is thought to be the cause
of the low densities and lack of recovery of the Venezuelan queen conch
population (CITES 2003). Connectivity models show Venezuela is largely
self-recruiting (Vaz et al. 2022); thus, queen conch in Venezuelan
waters must maintain relatively high adult densities to support
recruitment and population growth. Therefore, without adequate
enforcement of current regulations prohibiting the harvest of the local
queen conch population, which are already depleted and unlikely to be
successfully reproducing, densities will likely continue to decline
into the future.
Western Caribbean (Mexico, Belize, Honduras)
The jurisdictions in the western Caribbean have a history of
industrial-scale exploitation of queen conch. In Mexico and Belize, the
queen conch fisheries grew rapidly during the 1970s, which was followed
by subsequent declines in queen conch population and densities (CFMC
and CFRAMP 1999). In Mexico, the government responded to these declines
by implementing temporary and permanent fishery closures in various
areas in the 1990s (CITES 2012). Despite these closures and the more
recent implementation of size limits, closed seasons, and quotas,
Mexico's queen conch population has largely failed (CITES 2012).
Density surveys conducted in 2009 show a population that is unlikely to
be reproductively viable (De Jes[uacute]s-Navarrete and Valencia-
Hern[aacute]ndez 2013). While Mexico reported in 2018 that there have
been no legal exports of wild queen conch from Mexico during the
previous 7 years (CITES 2018), the FAO data show queen conch exports
from Mexico increasing from 204 mt in 2003 to 623 mt in 2018 (see S2 in
Horn et al. 2022). Given that harvest and export of the already
depleted queen conch population in Mexico is still occurring, existing
regulatory measures are inadequate to protect the species from
overutilization and further decline. Additionally, illegal fishing of
queen conch at both the Chinchorro and the Cozumel Banks and at
Alacranes Reef is thought to be a significant factor inhibiting
recovery (CITES 2012).
In Belize, the heavy exploitation of queen conch almost led to a
stock collapse in 1996 (CITES 2003). In response, the government
prohibited the selling of diced conch (Government of Belize 2013),
instituted minimum shell length (178 mm) and clean meat weight
requirements (85 g) to prevent the harvest of immature conch,
prohibited harvest by SCUBA gear, and established a TAC limit based on
biennial surveys (Gongora et al. 2020). While the biennial surveys to
determine TAC show relative stability in queen conch size classes over
several years, there is evidence of potential overutilization. For
example, Foley and Takahashi (2017) found that only 50 percent of
female conch were mature at 199 g (clean market meat), which is
significantly higher than the current minimum 85 g weight requirement,
indicating that this requirement is too low to protect immature conch.
In addition, Tewfik et al. (2019) documented a significant 15-
[[Page 55221]]
year decline in the mean shell length of adult and sub-adult queen
conch at Glover's Atoll, likely due to the selective harvest of conch
with a certain shell length size. This decline in the size distribution
may impact productivity because smaller adults tend to have lower
mating frequencies and smaller gonads (Tewfik et al. 2019), thereby
leading to a decline in overall reproductive output.
Tewfik et al. (2019) found evidence that indicates Belize's minimum
shell length size (178 mm) and market clean meat (85 g) regulations are
inadequate to protect juveniles from harvest. Tewfik et al. (2019) also
found a significant amount of immature conch with shell length sizes
over 178 mm and suggest lip thickness should be used as a proxy for
maturity, rather than shell length. Based on surveys of queen conch at
Glover's Atoll, Tewfik et al. (2019) calculated a threshold for the
size at 50 percent maturity to be a 10 mm thick shell lip and an
associated 192 g market clean meat. However, in Belize, queen conch are
not required to be landed intact with the shell. Because most conch
meat is removed at sea and the shell discarded, it is the minimum shell
size regulations are difficult to enforce and meat weight requirements
have diminished value in protecting undersized conch from harvest.
Based on the preceding, existing regulations are likely inadequate to
protect immature queen conch from harvest and may lead to a decline in
recruitment and growth in the future. In fact, the fishing of immature
queen conch has been confirmed directly by fishermen and fishery
managers, who note that imposing a lip thickness requirement would
significantly affect their landings as ``the majority of conch that is
fished are juveniles'' (Arzu 2019; FAO Western Central Atlantic Fishery
Commission 2020). In addition, a study conducted by Huitric (2005)
presented a historical review of conch fisheries and sequential
exploitation. The overall objective of this study was to analyze how
Belize's conch fisheries have developed and responded to changes in
resource abundance. Huitric (2005) suggests that the use of new
technology over time and space (by increasing the area of the fishing
grounds), together with fossil fuel dependence and fuel cost, have
sustained yields at the expense of depleted stocks, preventing learning
about resource and ecosystem dynamics, and removing incentives to
change fishing behavior and regulation.
Belize has established a network of marine reserves along the
Belize Barrier Reef and two offshore atolls that are divided up into
zones of varying levels of protection; however, enforcement of
protected areas is limited. For example, long-term declines of
reproductively active adult conch have been reported within the Port
Honduras Marine Reserve (PHMR) in southern Belize, a no-take zone for
queen conch. In fact, densities of conch have been continuously
declining since 2009, falling below 88 conch/ha by 2013, and decreasing
further to less than 56 conch/ha in 2014 (Foley 2016, unpublished cited
in Foley and Takahashi 2017). There have also been reports of illegal
fishing near Belize's border with Guatemala as well as reports of
Honduras fishermen illegally selling seafood products from Belize (Arzu
2019). In 2017, the Belize Fisheries Department reported confiscating
around 4.1 mt of queen conch meat that was harvested out of season (San
Pedro Sun 2018). The existing regulations appear adequate to maintain a
conch fishery in the short-term because there at least some large
mature conch that are protected from fishing located below the depths
usually accessed by free-diving (Tewfik et al. 2019;
Singh[hyphen]Renton et al. 2006). But the existing regulations will
likely be inadequate to prevent overutilization of the species in the
future, in light of the evidence of significant harvesting of immature
queen conch, the decreasing size of adult queen conch in the
population, ongoing reports of IUU fishing, and lack of enforcement.
Further, Tewfik et al. (2019) found that the deep water sites (i.e.,
fore-reef sites at Glovers Atoll), which are generally protected from
fishing due to their location, displayed the lowest overall density
(14-4 conch/ha) and were dominated by significantly older individuals
(lip thickness >20 mm) that have lower fecundity.
Honduras is one of the largest producers of queen conch meat, with
some population monitoring and evidence of general compliance with
existing regulations; however, there is also substantial evidence of
IUU fishing. In 1996, visual surveys resulted in an overall juvenile
and adult density of 14.6 conch/ha (Tewfik et al. 1998b). These low
densities were attributed to intensive exploitation that had taken
place over the previous decades (CITES 2012). However, the most recent
survey available conducted in 2011 reported overall conch densities
that should be able to sustain successful reproductive activity at two
of the three major banks: 134 conch/ha at Roselind; 196 conch/ha at
Oneida; and 93 conch/ha at Gorda Banks (Regalado 2012). However, no age
structure data was provided with this survey, and therefore the SRT was
unable to determine what proportion of the population surveyed are
adult queen conch. However, the densities increased with depth, which
is most likely the result of fishing effort focused in shallow areas
(Regalado 2012). In the early 2000s, there was also evidence that a
significant portion of the queen conch meat landed in and exported from
Honduras was fished illegally from neighboring jurisdictions. In
particular, concerns were raised about a period when Jamaica's fishery
at Pedro Bank was closed (2000-2002), which led to an increase in
illegal fishing by foreign vessels (including Honduran vessels) and
coincided with an increase in queen conch meat exports from Honduras
(CITES 2003; CITES 2012). From 1999 to 2001, Honduras almost doubled
its queen conch production, elevating concerns about IUU fishing (FAO
2016). Honduras, in addition to other jurisdictions, was also
implicated in unlawful queen conch exports that were confiscated in
2008 during the Operation Shell Game investigation (U.S. House,
Committee on Natural Resources, 2008). Illegal fishing has been
connected to illegal drug trafficking, increasing the complexity of the
issue for fisheries managers and the enforcement challenges (FAO 2016;
<a href="http://canadianbusiness.com">canadianbusiness.com</a>, Illegal trade: raiders of the lost conch, April
28, 2008).
Due to the high amount of exports, lack of landings records,
evidence of illegal activity, and low population densities, Honduras
was placed under a CITES trade suspension in 2003, and the Honduran
government declared a moratorium on conch fishing from 2003 to 2006.
From 2006 to 2012, export quotas were set annually for queen conch meat
that was taken during scientific surveys (CITES 2012; Regalado 2012).
However, based on surveys in 2009-2011 at the three main queen conch
fishing banks (Regalado 2012), the mean queen conch landings from 2010
through 2018 represented about 12.3 percent of the standing stock, or
more than 50 percent above the recommendation to fish at 8 percent of
standing stock, indicating that quotas are being set too high to
sustain fishing of these queen conch populations (Horn et al. 2022). In
2012, Honduras lost a substantial portion of its conch fishing grounds
to Nicaragua in a marine dispute resolution (Grossman 2013). Subsequent
to that determination, Honduras terminated its queen conch research
program and temporarily ceased generating scientific reports to inform
the annual quota allocation.
In 2017, Honduras developed and adopted a formal fishery management
[[Page 55222]]
plan aimed at establishing legal and technical regulations contributing
to the sustainable use of its queen conch populations. Regulations
implemented in the plan established a quota of 310 mt of 100 percent
clean conch meat to be distributed among 11 industrial fishing vessels.
In 2018 and 2019, the total quota increased to 416 mt and was allocated
among 13 vessels. Each vessel must carry a satellite monitoring and
tracking system during operations and carry one inspector onboard.
Minimum size limits were also established at 210 mm shell length, 18 mm
shell lip thickness, and a minimum meat weight of 125 g. As previously
noted, minimum shell length and meat weight regulations are unreliable
since large juveniles can have larger shells and more meat than mature
adults. The minimum shell lip thickness of 18 mm likely prohibits
immature queen conch from harvest. However, shells are commonly
discarded at sea, as the existing regulations do not require queen
conch to be landed with the shell intact, which makes it difficult to
ensure compliance and enforcement of most size-based regulations. The
most recent data (for 2018-2019) show that approximately 416 mt of
clean conch meat was landed (Ortiz-Lobo 2019). However, 0.6 mt of conch
meat was seized by the Honduran Navy from an unauthorized vessel in
November 2018 (Ortiz-Lobo 2019), indicating IUU fishing is still a
problem. In addition, fishermen, who agreed to conduct population
abundance and density surveys as part of a condition to fish for queen
conch under CITES, reversed their decision (Ortiz-Lobo 2019), and
abundance surveys from which harvest quotas are established have not
been conducted since 2011. The evidence of IUU fishing and the failure
to conduct required stock surveys, while increasing export quotas,
suggests the existing regulatory measures, including the current
allowable quota, are likely inadequate to prevent further declines of
the Honduran population of queen conch in the future.
Windward Islands (Barbados, Dominica, Grenada, St. Lucia, St. Vincent
and the Grenadines, Trinidad and Tobago)
In the Windward Islands, queen conch populations appear to be
following the same trend as the Leeward Islands, likely due to Allee
effects and lack of self-recruitment. Connectivity models (Vaz et al.
2022) demonstrate that queen conch in the southern Windward Islands
(i.e., Barbados, Grenada, and Trinidad and Tobago) are mostly self-
recruiting with larvae hatching and being retained locally; however, it
is likely that little to no recruitment is occurring due to the
relatively low adult queen conch densities observed throughout the
Windward Islands. These low conch densities appear to be the result of
overexploitation through sustained and unregulated or inadequately
regulated queen conch fishing over the last several decades.
In Barbados and Trinidad and Tobago, there is no management of the
queen conch fishery or regulations pertaining specifically to queen
conch harvests or sales. While there are no queen conch surveys or
assessment for Trinidad and Tobago, declines in abundance were noted as
early as the 1970s and 1980s (Georges et al. 2010; van Bochove et al.
2009; Luckhust and Marshalleck 2004; Lovelace 2002; Brownell and
Stevely 1981; Percharde 1968). In a 2010 technical report, 71 percent
of fishers interviewed reported declines in queen conch abundance
(Georges et al. 2010). Queen conch have been overfished and considered
depleted in Trinidad and Tobago since the 1990s (CITES 2012). In
Barbados, the queen conch catch is mainly comprised of immature
individuals, with an estimate as high as 96 percent (Oxenford and
Willoughby 2013), indicating highly unsustainable fishing of queen
conch. While there is limited information available on queen conch in
Dominica, the Significant Trade Review undertaken in 1995 resulted in a
CITES suspension of exports from Dominica (Theile 2001).
Grenada has been under a CITES trade suspension since May 2006 due
to failure to implement Article IV of the Convention, which requires
that the scientific authority of the state has advised that exports
will not be detrimental to the survival of the species (a determination
known as a `non-detriment finding'). During this trade suspension,
Grenada has continued to export conch to Trinidad and Tobago, and
Martinique (exporting 249 mt from 2007-2018; see S2 in Horn et al.
2022). However, Grenada recently indicated that it would be working
towards a regional action plan for queen conch in an effort to overcome
the CITES trade suspension (Blue BioTrade Opportunities in the
Caribbean, March 22-23, 2021).
St. Vincent and the Grenadines have regulations in place intended
to ensure sustainable conch fishing (FAO 2016). However, regulations
have not been updated since they were established in 1987 (Isaacs
2014), and queen conch density has continued to decline since the late
1970s, with estimates of 73 to 78 percent declines, depending on depth
area, from 2013 to 2016 (Rodriguez and Fanning 2018). Overall, adult
conch density estimates (10.4 conch/ha) are well below the minimum
adult density required to support any reproductive activity. Divers
have begun using SCUBA gear to reach deep waters as populations have
become depleted (CITES 2012). Current regulations prohibit the harvest
of queen conch with a shell length less than 18 cm, or without a flared
lip, or animals whose total meat weighs less than 225 g. Seasonal
closures have not been established and divers fish conch year round
(Rodriguez and Fanning 2018; CITES 2012). An export quota was
established, based on one of the highest export years recorded in 2002;
however, there appears to be no scientific basis for the establishment
of the export quota (CITES 2012). In fact, the high level of exports
that occurred in 2002 and 2004, was stated to be ``influenced by market
forces rather than stock abundance'' (Management Authority of St.
Vincent and the Grenadines in litt. to CITES Secretariat, 2004, as
cited in CITES 2012). The best available information indicates that
existing regulatory measures are inadequate to protect spawning adults,
as there is no seasonal closure, and deep water locations are being
fished with SCUBA gear. The existing regulations do not include a
minimum lip thickness requirement, a more reliable indicator of
maturity, to prevent harvest of immature conch and protect spawning.
Furthermore, because the existing quota system does not appear to be
based on population assessments or surveys, effective monitoring of the
fishery is lacking, which has likely contributed to the continued
depletion of the queen conch population.
In St. Lucia, the Department of Fisheries implemented regulations
in 1996 that prohibit the harvest of queen conch with a shell length
less than 18 cm, or without a flared lip, or animals whose total meat
weighs less than 280 g without digestive gland (Hubert-Medar and Peter
2012). Conch are harvested in St. Lucia mainly with SCUBA gear. There
are no lip thickness regulations to prohibit the harvest of juveniles,
and as previously described, shell length and flared lip are not
reliable indicators for maturity in conch. In addition, although the
Department of Fisheries requires queen conch to be landed whole in the
shell, it appears the majority of conch meat is extracted at sea and
the shell discarded (Williams-Peter 2021), making the shell length,
flared lip and meat weight requirements ineffective mechanisms for
protecting the fishery. Queen conch are also fished year round;
[[Page 55223]]
thus, fishing of spawning adults during their reproductive season is
likely occurring (Williams-Peter 2021). Information on stocks is still
scarce, especially information on density, abundance, and distribution
(Williams-Peter 2021). However, CPUE and landings data (1996-2007)
shows that stock have been in a steady decline (Williams-Peter, 2021;
Hubert-Medar and Peter 2012) indicating inadequate regulatory controls.
The best available information suggests that most jurisdictions
within the Windward Islands use inadequate proxy measures (i.e., shell
length, flared lip, and meat weight) to indicate maturity, allowing for
immature conch to be harvested. In addition, there is a general lack
monitoring of these fisheries to form the basis for their fishing
quotas, poor enforcement, and evidence IUU fishing. The connectivity
model (Vaz et al. 2022) indicates a strong reliance on self-recruitment
for these jurisdictions (although there is some exchange within
islands), with many of these jurisdictions acting as sources rather
than sinks for queen conch larva. Thus, it is likely that queen conch
throughout the Windward Islands will continue to decline due to
overutilization and the inadequacy of the existing regulatory measures
to address this threat.
Summary of Findings
Given the ongoing demand for queen conch, the lack of compliance
with and enforcement of existing regulatory measures, size-based
regulations that do not effectively protect juveniles from harvest, and
continued illegal fishing and international trade of the species,
combined with the observed low densities and declining trends in most
of the queen conch populations, the best available scientific and
commercial information indicates that existing regulatory mechanisms
are generally inadequate to control the threat of harvest and
overutilization of queen conch throughout its range. Our review of
minimum meat weight, shell length, and flared lip regulations indicates
that immature queen conch are being legally harvested in 20
jurisdictions, which is partially responsible for observed low
densities and declining populations. Shell lip thickness is considered
the most effective criterion for preventing the legal harvest of
immature queen conch (Appeldoorn 1994; Clerveaux et al. 2005; Cala et
al. 2013; Stoner et al. 2012; Foley and Takahashi 2017), while flared
shell lip and minimum shell length requirements do not guarantee sexual
maturity. Furthermore, there is general agreement among fisheries
managers that no individuals should be harvested before they have had
the opportunity to reproduce during at least one season (Stoner et al.
2012). Thus, the intent of the minimum size regulations is to protect
individuals until they have had the chance to reproduce at least once,
assuming that this will return a sustainable supply of new recruits
into the population. Nevertheless, only six jurisdictions (i.e.,
Colombia, Puerto Rico, Nicaragua, U.S. Virgin Islands, Cuba, and
Honduras) have minimum shell lip thickness regulations, but only
Honduras has a minimum shell lip thickness of at least 18 mm, which is
likely the most effective criteria for prohibiting the harvest of
immature conch; the other five jurisdictions require a minimum lip
thickness that may not ensure maturity (i.e., 5 mm, Colombia; 9.5 mm,
Puerto Rico; 9.5 mm, Nicaragua; and 10 mm, Cuba). While historical
studies report that some queen conch mature with relatively thin lips
(less than 7 mm) (Egan 1985, Appeldoorn 1988), more recent studies
indicate that maturation occurs later, at larger sizes, and differs by
gender (Doerr and Hill 2018). Several more recent studies indicate that
shell lip thickness values at maturity for queen conch range from 17.5
to 26.2 mm for females, and 13 to 24 mm for males (Avila-Poveda and
Barqueiro-Cardenas 2006; Aldana-Aranda and Frenkiel 2007; Bissada 2011;
Stoner et al. 2012). These studies have advocated for increases in the
minimum shell lip thickness for legal harvest. Avila-Poveda & Baqueiro-
C[aacute]rdenas (2006) suggests a minimum up to 13.5 mm by and Stoner
et al. (2012) suggests 15 mm. While, we recognize that the
relationships between shell lip thickness, age, and maturity vary
geographically, the best available information demonstrates that the
value established for minimum shell lip thickness by most jurisdictions
is inadequate to prevent immature conch from being harvested. In
addition, the majority of queen conch fisheries (except St. Lucia and
the U.S. Virgin Islands) do not have requirements to land queen conch
in the shell. Queen conch meat is typically removed and shell is
discarded at sea, which undermines enforcement and compliance with
regulations for a minimum shell length, shell lip thickness, and flared
shell lip. Furthermore, most jurisdictions require a minimum meat
weights (125 g to 280 g); however, meat weight is more applicable to
catch data, and generally does not constitute a reliable indicator of
queen conch maturity (FAO 2017). In addition, 15 jurisdictions do not
have regulations that include a seasonal closure, which is essential to
prevent the harvest of spawning adults. Similarly, 21 jurisdictions do
not have regulations that prohibit the use of SCUBA gear, which could
aid in protecting putative deep-water populations. Only a fraction of
the jurisdictions (i.e., Belize, The Bahamas, Jamaica, Nicaragua, and
Colombia) that have conch fisheries are conducting periodic surveys to
gather relevant information on the status of their queen conch
populations to inform their national management (e.g., TACs). Available
landings data indicate that substantial commercial harvest has led to
declines in many queen conch populations to the point where
reproductive activity and recruitment has been significantly impacted,
particularly throughout the eastern, southern, and northern Caribbean
region. Furthermore, several jurisdictions (e.g. Curacao and Trinidad
and Tobago) have no regulations despite having queen conch fisheries
(see S1 in Horn et al. 2022). Finally, Aruba (closed 1987), Bermuda
(closed 1978), Costa Rica (closed 1989), Florida (closed 1975), Panama
(closed 2004), and Venezuela (closed 2000) have completely closed their
respective queen conch fisheries. We conclude that fishery closures are
likely adequate, if enforced, to prevent further overutilization.
However, based on the longevity of the closures, and the lack of
recovery observed in each population, it is likely additional measures
will be necessary to restore those queen conch populations.
In summation, in some jurisdictions, regulatory controls are non-
existent. In other jurisdictions, fishery management regulations aimed
at controlling commercial harvest have fallen short of their goals,
largely due to a lack of population surveys, assessments, and
monitoring, and a reliance on minimum size-based regulations that
likely do not prevent the harvest of immature conch or protect spawning
stocks. In addition, poor enforcement and compliance with existing
regulations combined with significant IUU fishing has greatly reduced
the effectiveness of existing regulations. Based on the above, we
conclude that the best available information demonstrates that the
existing regulatory mechanisms throughout the range of the species are
inadequate to achieve their purpose of protecting the queen conch from
unsustainable harvest and continued populations decline.
[[Page 55224]]
Other Natural and Manmade Factors Affecting Its Continued Existence
Direct Impacts to Queen Conch From Climate Change
Queen conch reproduction is dependent on water temperature (Aladana
Aranda et al. 2014; Randall 1964), and therefore alteration to water
temperature regimes may limit the window for successful reproduction.
An increase in mean sea-surface temperatures may have direct effects on
the timing and length of the reproductive season for queen conch and
ultimately decrease reproductive output during peak spawning periods
(Appeldoorn et al. 2011; Randall 1964). Queen conch reproduction begins
at around 26-27 [deg]C. Aldana-Aranda and Manzano (2017) observed that
nearly all reproduction ceased when temperatures reached 31 [deg]C.
Early life history stages of queen conch are particularly sensitive to
ocean temperature (Brierley and Kingsford 2009; Byrne et al. 2011;
Harley et al. 2006), and rising water temperatures may have a direct
impact on larval and egg development (Aldana-Aranda and Manzano 2017;
Ch[aacute]vez Villegas et al. 2017; Boettcher et al. 2003). Aldana-
Aranda and Manzano (2017) tested the influence of climate change on
queen conch, larval development, growth, survival rate, and
calcification by exposing egg masses and larvae to increased
temperatures (28, 28.5, 29, 29.5 and 30 [deg]C, for 30 days. Queen
conch egg masses exposed to water temperatures greater than 30 [deg]C
resulted in the highest larval growth rate, but also higher larval
mortality (76 percent; Aldana-Aranda and Manzano 2017). This study
found no link between elevated water temperatures and the calcification
process in queen conch larvae. Furthermore, heat stress can induce
premature metamorphosis of queen conch leading to developmental
abnormalities and lower survival (Boettcher et al. 2003). Higher
temperatures also accelerate growth rates and decrease the amount of
time queen conch spend in vulnerable early stages. For example, faster
growth of juvenile queen conch offers earlier protection from predators
and shortens the time to reach sexual maturity. While growth may be
optimized at higher temperatures up to a certain point, the evidence to
date suggests that warming ocean conditions will also lead to higher
queen conch mortality rates for early life stages and possible
disruption of the shell biomineralization process (Aldana-Aranda and
Manzano 2017; Ch[aacute]vez Villegas et al. 2017). In addition, other
studies have indicated that queen conch veligers developed normally at
28 [deg]C, decrease growth at 24 [deg]C and have 100 percent mortality
at 32 [deg]C (Glazer pers. comm, as cited in Davis 2000; Aldana Aranda
et al. 1989; Aldana Aranda and Torrentera 1987.). However, Davis (2000)
found that a temperature of 32 [deg]C provided conditions for fast
growth and high survival of veligers, but also noted this temperature
is probably near the upper physiological tolerance for these veligers.
These findings suggest that future water temperatures in the Caribbean
Sea are likely to impact survival rates of queen conch during its early
life stages.
Climate change will also adversely impact the Caribbean region
through ocean acidification, which affects the calcification process of
organisms with calcareous structures, like the shells of queen conch.
Ocean acidification impedes calcareous shell formation, and thereby
impacts shell development (Aldana-Aranda and Manzano 2017; Parker et
al. 2013). Many mollusks, like queen conch, deposit shells made from
calcium carbonate (CaCO<INF>3;</INF> in the form of aragonite and high-
magnesium calcite) and these shells play a vital role in protection
from predators, parasites, and unfavorable environmental conditions.
Low pH is known to have a strong negative impact on larval development
in mollusks, like queen conch, and the very thin shells of queen conch
veligers may be especially vulnerable (Chavez-Villegas et al. 2017).
The absorption of CO<INF>2</INF> into the surface ocean has led to
a global decline in mean pH levels of more than 0.1 units compared with
pre-industrial levels (Raven et al. 2005, Parker et al. 2013). A
further 0.3 to 0.4 unit decline is expected over this century as the
partial pressure of CO<INF>2</INF> (pCO<INF>2</INF>) reaches 800 ppm
(Raven et al. 2005; Feely et al. 2004). At the same time there will be
a reduction in the concentration of carbonate ions (CO<INF>3</INF>\-
2\), which will lower the CaCO<INF>3</INF> saturation state in
seawater, making it less available to organisms that use
CaCO<INF>3</INF> for shells development (Cooley et al. 2009; as cited
in Parker et al. 2013). Ocean acidification impacts to larval queen
conch could have major impacts on recruitment to the adult age class,
including reproductive populations, throughout the species'
distribution (Stoner et al. 2021). Whether the impacts of ocean
acidification persist over multiple generations and at large enough
spatial scales to affect the long-term viability of queen conch
populations remains uncertain (Aldana-Aranda and Manzano 2017; Gazeau
et al. 2013). While changes to ocean pH will likely upset the shell
biomineralization processes, and challenge metabolic processes and
energetic partitioning, acidic ocean conditions can be patchy in space
and time and may develop slowly (Aldana-Aranda and Manzano 2017).
Research conducted by Aldana-Aranda and Manzano (2017) observed that
acidification conditions produced a 50 percent decrease in aragonite in
queen conch larval shell calcification at pH 7.6 and 31 [deg]C (see
Figure 21 in Horn et al. 2022). As previously mentioned, aragonite and
high-magnesium calcite are the primary ingredients in queen conch shell
formation. Uncertainty with regard to the queen conch's ability to
adapt to predicted changing climate conditions, the potential costs of
those adaptations, and the projections of future carbon dioxide
emissions make it difficult to assess the severity and magnitude of
this threat to the species. Recent studies and reviews have stressed
the importance of conducting multi-stressor (e.g., elevated water
temperature and ocean acidity), multi-generational, and multi-predicted
scenario experiments using animals from different areas in order to
better understand the impacts of climate change on mollusks at species-
wide levels (Aldana-Aranda and Manzano 2017; Parker et al. 2013).
Indirect Impacts to Queen Conch From Climate Change
Queen conch nursery habitat includes shallow and sheltered back
reef areas that contain moderate amounts of seagrass. These areas are
characterized by strong tidal currents and frequent exchange of clear
seawater (Stoner et al. 1996). Sea level rise, erosion, sea surface
temperatures, eutrophication, turbidity, siltation, and severity of
hurricanes and tropical storms resulting from climate change can have
both short- and long-term impacts on the water quality and health of
seagrass meadows (Boman et al. 2019; Cullen-Unsworth et al. 2014; Grech
et al. 2012; Burkholder et al. 2007; Orth et al. 2006; Duarte 2002;
Short and Neckles 1999). Depending on the frequency, severity, and
scale of climate change-induced conditions, seagrass meadow biomass may
decrease at local and over larger scales, reducing conch larvae
encounter rates with appropriate queen conch veliger settlement cues
(i.e., Thalassia testudinum detritus and associated epiphytes; Davis
and Stoner 1994). In addition, high water temperatures (greater than 30
[deg]C) in the shallow flats where queen conch nurseries occur can
result in low oxygen concentrations, which would reduce queen conch
growth and may lead to maturation at smaller than normal length,
thereby impacting reproductive output (Stoner
[[Page 55225]]
et al. 2021). Juvenile queen conch may experience lower growth and
higher mortality rates if they have limited access to adequate food
sources and shelter from predators, which are also provided by seagrass
meadow communities (Appeldoorn and Baker 2013). Deposits of fine
sediment or sediment with high organic content in a wider variety of
habitats that adults depend upon (e.g., algal plains, coarse sand,
coral rubble, and seagrass meadows) could smother the algae queen conch
graze on, thus limiting the nutritional value, and making these
habitats unsuitable (Appeldoorn and Baker 2013).
Queen conch are described as stenohaline (Stoner 2003), meaning
they tolerate a narrow range of salinities (approximately 34-36 ppt).
The species' ability to adapt to short- or long-term intrusions of
lower salinity water is uncertain; however, in at least one
groundwater-fed coastal area on the Yucatan Peninsula, queen conch
movement and growth was not different from core habitat areas with more
stable salinity and temperature signatures (Dujon et al. 2019;
Stieglitz et al. 2020). Hypoxic or anoxic conditions may also affect
the movement of juvenile queen conch (Dujon et al. 2019), which could
make them more vulnerable to predation. Changing climate may have
subtler effects that could impact tidal flow, circulation patterns, the
frequency and intensity of storm events, and larger scale current
patterns (Franco et al. 2020; van Gennip et al. 2017). Changes in tidal
flow and current patterns could alter the rate and condition of larval
dispersal and the cycle of source and sink dynamics of queen conch
populations throughout the Caribbean region. Changes in circulation
patterns within the Caribbean Sea would have significant implications
for the species.
Summary of Findings
The most significant impacts to queen conch resulting from climate
change are increased ocean temperature, ocean acidification, and
possible changes in Caribbean circulation patterns. According to
several studies, previously discussed, an increase in CO<INF>2</INF>
expected by the year 2100 is likely to negatively impact shell
formation, since water conditions will be more acidic and potentially
dissolve the shells of many mollusks. These studies have also suggested
that decreases in aragonite and larval shell calcification occur at a
pH 7.6-7.7, which is projected to occur by 2100 under the very high
greenhouse gas emissions scenario (SSP5-8.5; IPCC 2021). These changes
in water parameters are likely to result in significantly weaker and
thinner shells, which may increase predation rates, thereby
contributing to another source of mortality for the species in the
foreseeable future. Similarly, changes to other water parameters (e.g.,
salinity and dissolved oxygen) outside the range of those typically
experienced by queen conch can impact their growth and survival and
have negative consequences on the seagrass habitat upon which they
depend.
The most recent Intergovernmental Panel on Climate Change (IPCC)
projections indicate that mean sea surface temperature will warm by
3.55 [deg]C by 2100, with the increase in sea surface temperature
ranging from 2.45 [deg]C to 4.85 [deg]C. The available information
indicates that the Caribbean Sea will follow the global mean
temperature (IPCC 2021; Figure SPM.5). The temperature of the Caribbean
Sea has warmed to approximately 28 [deg]C at present (Bove et al.
2022). Thus, based on the IPCC projections for mean sea surface
temperature, it appears that water temperature may increase by
approximately 3.55 [deg]C suggesting that Caribbean Sea surface
temperatures will exceed 31 [deg]C under scenario SSP5-8.5 by 2100
(IPCC 2021). A mean sea surface temperature in the Caribbean Sea in
excess of 31 [deg]C may have negative implications for early life
stages and queen conch reproduction. The impacts of acidification on
conch larvae could also have significant impacts on recruitment to the
adult class, including reproductive populations, throughout the
species' range. In addition, possible changes in Caribbean Sea
circulation patterns would have significant implications for queen
conch recruitment processes and reproduction, but the extent of the
impacts from changes in circulation patterns to queen conch is not well
understood. Even so, the information is alarming as it indicates that
the reproduction, growth, and survival of queen conch will likely be
impacted by climate change in the future.
Assessment of Extinction Risk
The ESA (section 3) defines an endangered species as ``any species
which is in danger of extinction throughout all or a significant
portion of its range.'' A threatened species is defined as ``any
species which is likely to become an endangered species within the
foreseeable future throughout all or a significant portion of its
range'' (16 U.S.C. 1532). Implementing regulations in place at the time
the status review was completed described the ``foreseeable future'' as
the extending only so far into the future as we can reasonably
determine that both the future threats and the species' responses to
those threats are likely. These regulations instructed us to describe
the foreseeable future on a case-by-case basis, using the best
available data and taking into account considerations such as the
species' life-history characteristics, threat-projection timeframes,
and environmental variability. The regulations also indicated that we
need not identify the foreseeable future in terms of a specific period
of time. Although these regulations were vacated on July 5, 2022, by
the United States District Court for the Northern District of
California and are thus no longer in effect, this approach for
determining the ``foreseeable future'' is consistent with NMFS's
longstanding interpretation of this term in use prior to the issuance
of these regulations in 2019 (see 84 FR 45020, August 27, 2019).
For the assessment of extinction risk for the queen conch, the
``foreseeable future'' was considered to extend out several decades
(approximately 30 years). Given the species' life history (i.e.,
density dependent reproduction and longevity estimated to be 30 years),
it would likely take more than several decades and multiple generations
for management actions to be reflected in population status. Similarly,
the impact of present threats to the species could be realized in the
form of noticeable population declines within this time frame, as
demonstrated in the available survey and fisheries data. We also
acknowledge that population recovery is likely dependent on when a
protective regulatory measure, such as a closure, is implemented and
the status of the population at the time of the closure. For example,
Florida, Bermuda, and Aruba prohibited all conch harvest in the mid
1980's (more than 35 years ago), yet their respective populations have
yet to recover. Other recovery efforts such as those in Cuba and on
Colombia's Serrana Bank were started earlier and recoveries occurred
over a shorter timeframe. In addition, in order to fully assess the
longer-term threats stemming from climate change and their impacts on
queen conch, we considered these threats over a time horizon that
extended out to 2100, which is the timeframe over which both climate
change threats and impacts to queen conch could be reasonably
determined, with increasing uncertainty in climate change projections
over that time period. Thus, while precise conditions during the year
2100 are not reasonably foreseeable, the general trend in conditions
during the period of time
[[Page 55226]]
from now to 2100 is reasonably foreseeable as a whole, although less so
through time.
Demographic Risk Analysis
In determining the extinction risk of a species, it is important to
consider not only the current and potential threats impacting the
species' status but also the species' demographic status and
vulnerability. A demographic risk analysis is an assessment of the
manifestation of past threats that have contributed to the species'
current status and informs the consi
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This is legal information, not legal advice. Laws vary by jurisdiction and change frequently. Always verify current law with official sources and consult a licensed attorney in your jurisdiction for advice on your specific situation.