Global patterns of illegal marine turtle exploitation

Abstract

Human exploitation of wildlife for food, medicine, curios, aphrodisiacs, and spiritual artifacts represents a mounting 21st‐century conservation challenge. Here, we provide the first global assessment of illegal marine turtle exploitation across multiple spatial scales (i.e., Regional Management Units [RMUs] and countries) by collating data from peer‐reviewed studies, grey literature, archived media reports, and online questionnaires of in‐country experts spanning the past three decades. Based on available information, we estimate that over 1.1 million marine turtles were exploited between 1990 and 2020 against existing laws prohibiting their use in 65 countries or territories and in 44 of the world's 58 marine turtle RMUs, with over 44,000 turtles exploited annually over the past decade. Exploitation across the 30‐year period primarily consisted of green (56%) and hawksbill (39%) turtles when identified by species, with hawksbills (67%) and greens (81%) comprising the majority of turtles exploited in the 1990s and 2000s, respectively, and both species accounting for similar levels of exploitation in the 2010s. Although there were no clear overarching trends in the magnitude or spatial patterns of exploitation across the three decades, there was a 28% decrease in reported exploitation from the 2000s to the 2010s. The 10 RMUs with the highest exploitation in the 2010s included seven green and three hawksbill turtle RMUs, with most reported exploitation occurring in RMUs that typically exhibit a low risk of population decline or loss of genetic diversity. Over the past decade, the number of RMUs with “moderate” or “high” exploitation impact scores decreased. Our assessment suggests that illegal exploitation appears to have declined over the past decade and, with some exceptions, is primarily occurring in large, stable, and genetically diverse marine turtle populations.

Full text

Glob Change Biol. 2022;28:6509–6523. wileyonlinelibrary.com/journal/gcb
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6509© 2022 John Wiley & Sons Ltd.
Received: 13 March 2022
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Revised: 12 July 2022
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Accepted: 12 July 2022
DOI : 10.1111/gcb .16378
RESEARCH ARTICLE
Global patterns of illegal marine turtle exploitation
Jesse F. Senko1 | Kayla M. Burgher2 | Maria del Mar Mancha- Cisneros3 |
Brendan J. Godley4 | Irene Kinan- Kelly5 | Trevor Fox2 | Frances Humber4,6 |
Volker Koch7 | Andrew T. Smith2 | Bryan P. Wallace8,9
Jesse Senko and K ayla Burgher sh ould be conside red joint firs t author.
1School for the Future of Innovation in
Society, Arizona St ate University, Tempe,
Arizona, USA
2School of Life Sciences, Arizona State
University, Tempe, Arizona, USA
3Scripps Institution of Oceanography,
University of California San Diego, La
Jolla, California, USA
4Marine Turtle Research Group, Centre for
Ecology and Conservation, University of
Exeter, Penr yn, UK
5NOAA Fisheries, Pacific Islands Regional
Office, Honolulu, Hawaii, USA
6Blue Ventures Conservation, London, UK
7Deutsche Gesellschaft für Internationale
Zusammenarbeit (GIZ) GmbH, Eschborn,
Germany
8Ecolibrium, Inc., Boulder, Colorado, USA
9Department of Ecology and Evolutionar y
Biology, University of Colorado Boulder,
Boulder, Colorado, USA
Correspondence
Jesse F. Senko, School for the Future
of Innovation in Society, Arizona State
University, Tempe, Arizona, USA.
Email: jesse.senko@asu.edu
Abstract
Human exploitation of wildlife for food, medicine, curios, aphrodisiacs, and spir-
itual artifacts represents a mounting 21st- century conservation challenge. Here,
we provide the first global assessment of illegal marine turtle exploitation across
multiple spatial scales (i.e., Regional Management Units [RMUs] and countries) by
collating data from peer- reviewed studies, grey literature, archived media reports,
and online questionnaires of in- country experts spanning the past three decades.
Based on available information, we estimate that over 1.1 million marine turtles
were exploited between 1990 and 2020 against existing laws prohibiting their use
in 65 countries or territories and in 44 of the world's 58 marine turtle RMUs, with
over 44,000 turtles exploited annually over the past decade. Exploitation across
the 30- year period primarily consisted of green (56%) and hawksbill (39%) turtles
when identified by species, with hawksbills (67%) and greens (81%) comprising the
majority of turtles exploited in the 1990s and 2000s, respectively, and both spe-
cies accounting for similar levels of exploitation in the 2010s. Although there were
no clear overarching trends in the magnitude or spatial patterns of exploitation
across the three decades, there was a 28% decrease in reported exploitation from
the 2000s to the 2010s. The 10 RMUs with the highest exploitation in the 2010s
included seven green and three hawksbill turtle RMUs, with most reported exploi-
tation occurring in RMUs that typically exhibit a low risk of population decline or
loss of genetic diversity. Over the past decade, the number of RMUs with “moder-
ate” or “high” exploitation impact scores decreased. Our assessment suggests that
illegal exploitation appears to have declined over the past decade and, with some
exceptions, is primarily occurring in large, stable, and genetically diverse marine
turtle populations.
KEYWORDS
aquatic bushmeat, black market, bycatch, directed take, fisheries, hunting, poaching, wildlife
trade

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1 | INTRODUC TION
Human exploitation (or take) of wildlife is a global threat to biodiver-
sity (Milner- Gulland & Bennett, 2003) and has received increased
at tentio n as a maj o r driv er in the em erge nce an d tra nsmi ssio n of zoo -
notic diseases, especially in light of the COVID- 19 pandemic (Volpato
et al., 2020). Illegal exploitation of wildlife for food, medicine, curios,
aphrodisiacs, spiritual artifacts, and other products is considered
one of the most valuable illicit markets in the world (Challender &
MacMillan, 2014; Haken, 2011), and has contributed to declines
in terrestrial and marine species worldwide (Bennett et al., 2002;
Groombridge & Luxmoore, 1989; Jerozolimski & Peres, 2003).
Despite increased awareness of the problem and several measures
to curb it, illegal trade of wildlife is growing and worth an estimated
20 billion USD worldwide (Challender & MacMillan, 20 14; Milliken
& Shaw, 2012; Underwood et al., 2013). High- profile examples in-
clude terrestrial species such as pangolins (Wu et al., 2004), ele-
phants (Underwood et al., 2013), tigers (Sharma et al., 2014), and
rhinoceroses (Ferreira et al., 2012), as well as marine species such
as sharks (Clarke et al., 2006), totoaba fish (Valenzuela- Quinonez
et al., 2015), cetaceans (Baker et al., 2010), humphead wrasse
(Donaldson & Sadovy, 2001), and marine turtles (Lopes et al., 2022;
Miller et al., 2019).
Marine turtles have been exploited for their meat, eggs, shell,
skin, and internal organs since at least 13,000 years before pres-
ent (Des Lauriers, 2006), and have historically been an import-
ant resource for coastal inhabitants worldwide (Barrios- Garrido
et al., 2018; Early- Capistrán et al., 2018, 2020; Frazier, 2003). By
the 18th century, marine turtles were exploited on commercial and
international scales for food and other products (Bell et al., 2007;
Cornelius et al., 2007; Early- Capistrán et al., 2018; Lam et al., 2012;
Limpus et al., 2008; McClenachan et al., 2006). During the mid-
19th century, industrial- scale trade in tortoiseshell products (i.e.,
carapace scutes from hawksbill turtles, Eretmochelys imbricata) was
well underway, primarily serving the Japanese market (Mortimer &
Donnelly, 2008). By the 20th century, rapidly increasing global de-
mand, globalization, and advances in transportation technology (e.g.,
roads connecting remote areas to urban center s) that increased con-
nectivity, paired with modern tools that facilitated capture (e.g., spear
guns, nylon nets/rope, aluminum and fiberglass boats, outboard
motors, and navigational systems), led to industrial- scale exploita-
tion that peaked with over 17,000 tons of marine turtle meat and
products (e.g., leather, tortoiseshell) taken in commercial fisheries
during the 1960s (FAO, 2011), including over 380,000 marine turtles
in a single year (1968) from Mexico alone (Cantú & Sánchez, 1999).
Recent estimates suggest that 9 million hawksbill turtles (or 60,000
turtles annually) were traded globally over an approximate 150- year
period (1844– 1992) for their shells (Miller et al., 2 019). Worldwide,
large- scale commercial exploitation has caused declines in several
marine turtle populations over the past half- century (Jackson, 1997;
Mortimer & Donnelly, 2008; Seminoff et al., 2015; Stoddart, 1980).
Today, while catch and utilization of marine turtles are still per-
mitted in some parts of the world (see Humber et al., 2014), most
countries have regulations that range from full protection (typically
with permitted incidental take in commercial fisheries) to regulated
take regimes (Brautigam & Eckert, 2006; CITES Secretariat, 2019 ;
Humber et al., 2014; Lopes et al., 2022; Maison et al., 2 010).
Exploitation of marine turtles continues for food, but in some cases
is driven by socio- cultural reasons that vary by region (Barrios-
Garrido et al., 2018; Lam et al., 2012; Mancini & Koch, 2009; von
Essen et al., 2 014). For example, marine turtle meat is considered a
delicacy in some regions of Latin America, such as northwest Mexico
(Mancini et al., 2011; Quiñones et al., 2017; Senko et al., 2010, 2014).
In China and Vietnam, whole stuffed turtles are seen as a symbol of
status and wealth (Hamann et al., 2006), while tortoiseshell products
are a traditional component of Japanese dress and remain highly val-
ued (Canin, 1991; Lam et al., 2012; Yifan, 2018). Turtle blood, fat,
and testicles are considered aphrodisiacs in East Africa, and live
turtles are offered to spirits in sacred religious ceremonies in the
Pacific Islands region (Álvarez- Varas et al., 2020; Rudrud, 2010;
Westerlaken, 2016).
Domestic and international conservation efforts, legislation,
and inter- governmental conservation structures (e.g., the U.S.
Endangered Species Act, Convention on International Trade in
Endangered Species of Wild Fauna and Flora [CITES], Convention
on the Conservation of Migratory Species of Wild Animals) enacted
over the past half- century appear to have had some success in re-
ducing impacts of exploitation and other threats, with global- scale
patterns that suggest more marine turtle populations appear to be
increasing versus decreasing (Mazaris et al., 20 17). Illegal exploita-
tion (sourced via direct capture or accidental and retained bycatch) is
still, however, likely prevalent in many regions worldwide. Moreover,
an evaluation of the conservation status of all marine turtle Regional
Management Units (RMUs) globally revealed that exploitation
was scored as the threat category with the second highest impact
to marine turtle RMUs, behind only fisheries bycatch (Wallace
et al., 2011). Assessing the magnitude, impacts, and trends of illegal
exploitation would therefore improve our understanding of the driv-
ers of population changes at both regional and global scales (Lopes
et al., 2022), while helping researchers and resource managers iden-
tify conservation priorities across geographies, species, and popula-
tions. Nevertheless, such an assessment has yet to be conducted on
a global scale.
Here, we assessed the global scope of illegal exploitation of ma-
rine turtles— defined broadly as the take or trade of turtles against
established law in the locality where the exploitation occurred— for
meat or products such as medicine, jewelry, leather, or taxidermies,
excluding eggs. We focused explicitly on illegal exploitation (here-
after referred to as ‘exploitation’), as legal fisheries generally have
specific management measures in place that limit species, number
of individuals, size classes, and/or seasons in which animals can be
taken (Brautigam & Eckert, 2006; CITES Secretariat, 2019; Humber
et al., 2014 ; Maison et al., 2010). Specifically, we analyzed data
from peer- reviewed studies, grey literature, archived media reports,
and online questionnaires with renowned marine turtle experts to
achieve the following research objectives:

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• Evaluate the magnitude and potential trends of marine turtle
exploitation globally, among countries, RMUs, and marine turtle
species between 1990 and 2020;
• Characterize geographic and temporal patterns in reported traf-
ficking of marine turtles and their products;
• Assess potential population- level impacts of exploitation on ma-
rine turtle RMUs.
We recognize the need to develop widely accepted, cultur-
ally appropriate terminology when describing the legality of ma-
rine turtle exploitation, as terms or activities can be difficult to
interpret depending on cultural norms generally and on relevant
rules of law in a specific country or territory (e.g., see Delisle
et al., 2018). We also recognize that there may be variations in
governance and application of the rule of law in a given country,
how laws exist in relation to cultural norms, and whether or not
they are ethical. Through out this pap er, we strived to handle thes e
topics with sensitivity and respect while fulfilling our overarching
goals, and recommend that standard, appropriate terminology and
assessment approaches be developed.
2 | METHODS
2.1 | Analytic goals
Our overarching goal was to assess the number of marine turtles
illegally exploited worldwide, along with trends in exploitation over
time, using data from peer- reviewed journal articles, grey literature,
archived media reports, and online questionnaires of in- country ex-
perts spanning the past three decades, 1990– 2020 (peer- reviewed,
grey, and media sources are herein referred to as “documented”
sources/dat a). Alth oug h not as rob ust as peer- r evi ewed article s, gr ey
literature, media reports, and online questionnaires are utilized in
addition to peer- reviewed articles to provide a more comprehensive
picture of illegal exploitation by filling knowledge gaps in areas lack-
ing peer- reviewed scientific research. For example, the news media
plays an important role in shaping public understanding of societal
problems, and analyses of media coverage have been conducted
on sharks (Shiffman et al., 2020), panthers (Jacobson et al., 2012),
grizzly bears (Hughes et al., 2020), and marine turtles (Santos &
Crowder, 2021). Similarly, grey literature such as NGO reports and
practitioner newsletters may constitute a significant proportion of
available data used to inform management decisions and can con-
tain vital information for conservation practitioners (Haddaway &
Bayliss, 2015). Finally, online questionaries are an increasingly wide-
spread tool, especially in light of the COVID- 19 pandemic, which
can provide ready access to large and diverse samples of people
(Wardropper et al., 2021).
In our analysis, we included tallies of whole turtles or products
that could be definitively linked to individual turtles (e.g., heads, tails,
or shells), but did not include products that could not be (e.g., eggs
or pieces of jewelry such as bracelets or earrings). To determine the
legality of marine turtle exploitation, we first reviewed and updated
the Humber et al. (2014) global review of legal marine turtle fisher-
ies. We then reviewed available information and/or responses from
in- country experts describing the legislation in each country with
documented exploitation. Marine turtles that were exploited under
cultural or indigenous take provisions (e.g., Australia), irrespective of
broader legislation in the country, were not included in the analysis.
All illegal exploitation of individual turtles, including non- permitted
take within legal take regimes, was included in our analysis. In cases
where legislation was unclear, or if we were unable to verify legis-
lation, the study and country were not included. Data from farming
or ranching operations were not included, while animals taken as
fisheries bycatch were only included if they were illegally retained.
We specif ical ly so ugh t data on the gl oba l mag n itude an d dis trib u-
tion (i.e., country or territory and RMU, the latter of which describes
spatially explicit marine turtle populations above the level of individ-
ual nesting rookeries and below the global species level, using “geo-
referenced available data on marine turtle biogeography— including
individual nesting sites, genetic stocks, and geographic distributions
based on monitoring research— to spatially integrate sufficient in-
formation to account for complexities in marine turtle population
structures”; Wallace, DiMatteo, et al., 2010 ) of exploitation, along
with data on which species were exploited. We classified turtles as
“trafficked” in the context of exploitation if they were known to have
moved across international bodies, if they were seized by authorities
during international transit, or if the data source otherwise stated
that the animal was internationally traded.
2.2 | Data compilation and considerations
We assessed the number of marine turtles exploited worldwide
from January 1, 1990 to December 31, 2019 (our last searches were
conducted on July 29, 2020) through a comprehensive collation and
synthesis of (1) publications in peer- reviewed journals; (2) technical
and agency documents, non- governmental organizations including
the IUCN Species Survival Commission's Marine Turtle Specialist
Group regional reports, TRAFFIC technical reports, and symposia
proceedings (collectively termed ‘grey literature’); (3) archived media
reports; and (4) an online questionnaire sent to in- countr y marine
turtle experts.
Our review of peer- reviewed and grey literature was conducted
by searching the following databases: ISI Web of Knowledge (i.e.,
Zoological Record, ISI Web of Science, and BIOSIS), Google Scholar,
Index of Online Theses and Dissertations, Directory of Open Access
Journals, and IUCN Library Databases (i.e., IUCN, CITES, TR AFFIC,
WIDECAST). We examined the first 150 document hits from each
Internet source for inclusion criteria. We used the following search
terms in our literature review (*indicates a wildcard), with terms
searched both independently and combined using AND/OR fea-
tures: (1) sea turtle utilize* OR utilis*; (2) marine turtle utilize* OR
utilis*; (3) sea turtle illegal; (4) marine turtle illegal; (5) sea turtle har-
vest*; (6) marine turtle harvest*; (7) sea turtle hunt*; (8) marine turtle

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hunt*; (9) sea turtle poach*; (10) marine turtle poach*; (11) sea turtle
trade*; (12) marine turtle trade*.
Our review of media articles was conducted by searching Arizona
State University's international newspaper database “Access World
News,” which houses more than 12,000 newspapers from all con-
tinents except Antarctica, for news articles (photographs without
associated articles were not included) using the aforementioned
search terms (results dated from January 1, 1990 to December 31,
2019, with the last search completed on July 29, 2020). All docu-
mented sources assessed were in English. The use of an international
news database allowed us to integrate news articles from a broad
array of sources, as opposed to using a handful of traditional outlets
(Santos & Crowder, 2021).
To ensure that we did not count duplicate incidents of marine
turtle exploitation, we compared the number of turtles exploited
as well as the year and location of exploitation for each new peer-
reviewed, grey, or media source against our existing database. If
the location and year of exploitation matched or were similar to
any entry in our existing database, we examined the details of the
information (e.g., number of turtles, people involved) to determine
whether the new source was a duplicate. If details of exploitation
matched an existing entry, the new source was not included.
While we treated each type of documented source equally when
counting illegally exploited turtles in our analysis, we calculated a
data quality score for each country to provide additional information
to interpret the robustness of results (sensu Wallace et al., 2011;
Dataset S1). To calculate the data quality scores, we assigned each
literature source a weight based on data quality as follows: media
articles were assigned a weight of 0.25; grey reports were assigned
a weight of 0.5 when a single method was used to assess illegal ex-
ploitation (e.g., only dumpsite surveys) and <50% of a country was
covered, and a weight of 0.75 when multiple methods were used to
assess illegal exploitation (e.g., dumpsite and market surveys) and
covered >50% of a country; peer- reviewed publications were as-
signed a weight of 0.75 when a single method was used to assess
illegal exploitation and <50% of a country was covered and a weight
of 1 when multiple methods were used to assess illegal exploitation
and covered >50% of a country. Then, for each country, a weighted
average was calculated to produce a data quality score by summing
each weight multiplied by the number of sources assigned that
weight, then dividing that by 1.5, the sum of the weights.
We used an online questionnaire of in- country marine turtle
experts to fill in data gaps (i.e., countries in which we were unable
to locate peer- reviewed, grey, or media sources), as illegal exploita-
tion is likely underreported due to its clandestine nature (Humber
et al., 20 11; Martin et al., 2012). Our voluntary confidential online
questionnaire (Google Forms) was distributed in English to known
in- country experts in marine turtle research and conservation living
in or known to work in target nations, and included experts identi-
fied in a database developed by Humber et al. (2014). Experts were
first identified if they had previously authored research on illegal
exploitation within a target nation. In cases where there were no
previous studies, we could not get in contact with the author(s) of
previous work, or there were no experts included in the Humber
et al. (2014) database, we used our collective experience in the field
of marine turtle research to identify in- country experts. We con-
tacted 104 experts from 80 countries via email and did not offer
compensation to participate. Experts were asked to provide esti-
mates on the levels of illegal exploitation over the past two decades
(i.e., 2000s and 2010s) for the applicable country or territory they
were requested to provide information on. Exploitation ranges pro-
vided in the questionnaire included: 0; 1– 100; >100– 1000; >1000–
10,000; >10,000– 25,000; >25,000– 50,000; >50,000– 75,000,
>75,000– 100,000; and >100,000 turtles per year, irrespective of
life stage or species. Respondents could also write a specific ex-
ploitation value other than the ranges provided. In cases where ex-
perts provided different exploitation ranges for the same country or
territor y, we used the lowest exploitation range to provide a conser-
vative estimate. Table S1 co nta ins the final explo ita tio n esti mates for
each country we received questionnaire responses for.
2.3 | Data analysis
Each documented data source was evaluated and filtered to only
include cases where exploitation was clearly illegal (i.e., performed
against existing laws prohibiting the use of marine turtles) during the
time it was documented. For countries whose legislation changed
during the study period (e.g., Vietnam), cases of exploitation were
only included when they occurred during active legislation prohibit-
ing marine turtle exploitation. From each documented source, we
extracted data for all relevant variables, which included the number
of tur tles of all lif e stages , dec ade of expl oit ation , loc atio n (i.e. , coun-
try or territory and RMU), species, and whether trafficking had oc-
curred. We only summarized data directly reported in data sources,
and did not calculate our own estimates or extrapolations, although
we did include annual estimates of exploitation reported in our col-
lated sources. When estimates were given as a total value for a range
of years that spanned multiple decades, we divided the number of
turtles exploited by the number of years provided to obtain a mean
estimate of annual exploitation so that the total number of turtles
exploited could be correctly apportioned by decade. Additionally,
when single annual estimates were given for a multiple- year study,
we multiplied the annual estimate by the length of the study in years
to find the total number of individuals. When no documented data
for a country or territory existed for a specific decade, we incorpo-
rated the exploitation estimates provided by the online question-
naire sent to in- country experts for that decade by multiplying the
median of the annual exploitation range by 10 to account for the en-
tire decade (see Dataset S1). We rounded all estimates of exploita-
tion that included questionnaire data to the nearest thousand when
possible to account for the imprecise nature of the ranges gathered
from the questionnaire data. We identified exploitation by species
and trafficked turtles only in cases where they were explicitly men-
tioned in the documented source. RMUs and trafficking status could
not be specified where the species and/or location of exploitation

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were not provided, which included all questionnaire data. When
RMUs overlapped, we split the number of turtles evenly between
the overlapping RMUs.
We created an exploitation impact score for each RMU to con-
textualize the magnitude and risk of exploitation relative to nesting
female population size, a standard population abundance metric for
marine turtles (NRC, 2010). This approach is akin to ‘bycatch impact
scores’ developed by Wallace et al. (2013) to estimate population-
level impacts of fisheries bycatch on marine turtle RMUs. Exploitation
impact scores were created for the entire study period along with
each respective decade. To calculate the exploitation impact scores
for the entire study period, we divided the average number of turtles
(of all life stages, males as well as females) exploited per decade by
estimates of abundance of nesting females obtained by calculating
the midpoints of population ranges provided in Wallace et al. (20 11).
For each individual decade, we divided the total number of turtles
exploited in that decade by the same estimates of nesting female
populations. An exploitation impact score, then, can be interpreted
as the approximate number of turtles exploited per nesting female
per decade. Therefore, when an exploitation impact score is greater
than 1, more turtles of all life stages and both sexes are being ex-
ploited within a 10- year period than there are adult females nesting
annually in the population. Additionally, we plotted the exploitation
impact scores for each RMU against the RMU risk scores (i.e., indices
of population viability that include criteria such as annual nesting
female abundance, trends, and genetic diversity) reported in Wallace
et al. (2011) to evaluate the impacts of exploitation relative to other
threats, as well as current population size and trajector y.
To provide appropriate context for strategic conservation pri-
ority setting, we developed conservation priorities for each RMU
that were based on our exploitation impact scores and risk scores
from Wallace et al. (2011). To create these conservation priorities,
we classified risk scores as either “low” or “high” according to the
method used in Wallace et al. (2011), and classified exploitation im-
pact scores as “low,” “medium,” or “high,” where “low” scores were
less than 0.5, “medium” scores were between 0.5 and 1, and “high”
scores were greater than 1. For the final conservation priority, we
reported both the level of risk and exploitation (i.e., low/medium/
high risk– low/medium/high exploitation).
Where data were available for multiple decades, we analyzed
decadal trends for the total number of turtles exploited, the spe-
cies exploited, RMUs (not including RMU exploitation impact score
since the trends follow those for the number of turtles exploited in
RMUs exactly), countries, and trafficking. In each trend analysis, the
trend was classified as “increasing,” “decreasing,” “no change,” or
“unclear.” An “unclear” classification was used for any case where
exploitation changed over time but did not consistently increase or
decrease across all decades for which data was available. We used
ArcDesktop 10 geographic information systems software to depict
the magnitude of exploitation for countries and territories using a
choropleth map for data from the literature review and graduated
symbols for data from the expert questionnaires. We also created
choropleth maps depicting exploitation for each species in their
respective RMUs. We used R 3.5.3 to plot and analyze all data (R
Core Team, 2019).
3 | RESULTS
3.1 | Overview of global exploitation
Summarizing 39 peer- reviewed publications, 42 grey reports, 82
media reports, and 46 expert questionnaire responses, an estimated
total of 1,128,000 marine turtles were reported to have been ex-
ploited between 1990 and 2020 across 65 countries or territories
and in 44 of the world's 58 marine turtle RMUs, representing approx-
imately 38,000 turtles annually across the three decades (Table S2;
Dataset S1). Documented data (1990– 2020) and questionnaire data
(2000– 2020) revealed changes in the magnitude of exploitation over
time, with 7875 turtles per year exploited in the 1990s, 61,000 tur-
tles per year exploited in the 2000s, and 44,000 turtles per year ex-
ploited in the 2010 s (Figure 1; Dataset S1), indicating a 28% decrease
in reported exploitation from the 2000s to the 2010s.
3.2 | Global trends in exploitation by species
Peer- reviewed, grey, and media sources documenting illegal marine
turtle exploitation between 1990 and 2020 reported 66% (131,519
turtles) of exploitation to the level of species (not including ques-
tionnaire data, which did not report on species). Across all decades,
green turtles (Chelonia mydas, 56%) were the most exploited species,
followed by hawksbills (Eretmochelys imbricata, 39%), loggerheads
(Caretta caretta, 3%), olive ridleys (Lepidochelys olivacea, 2%), leather-
backs (Dermochelys coriacea, <1%), and Kemp's ridleys (Lepidochelys
kempii, <1%; Dataset S1). Exploitation of flat back turtles (Natator de-
pressus) was not reported in the documented sources we reviewed.
On a finer decadal scale, 56% of exploited marine turtles were
reported by species in the 1990s, 95% in the 2000s, and 51% in the
2010s. Decadal analysis revealed changes over time in the propor-
tion of turtles of each species that were exploited. In the 1990s,
hawksbills were exploited most frequently (67%), followed by green
turtles (30%), loggerheads (2%), leatherbacks (0.5%), olive ridleys
(0.05%), and Kemp's ridleys (<0.001%; Figure 2a; Dataset S1). In the
2000s, the decade with the highest proportion of exploited turtles
with data reported to the level of species, green turtles were ex-
ploited the most (81%), followed by hawksbills (9%), loggerheads
(6%), olive ridleys (4%), leatherbacks (0.1%), and Kemp's ridleys
(0.002%; Figure 2b; Dataset S1). For the most recent decade (2010–
2020), hawksbills (54%) were the most abundant species, followed
by greens (45%), olive ridleys (0.4%), loggerheads (0.2%), and leath-
erbacks (0.04%; Figure 2c; Dataset S1). No Kemp's ridley turtles
were reported exploited in the 2010s. Of all species of marine tur-
tles reported to be exploited between 1990 and 2020, leatherbacks
were the only species to show a decline in numbers exploited across
all three decades.

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3.3 | Global distribution and trends in exploitation
by marine turtle RMUs
Sources that provided species- specific exploitation data revealed il-
legal marine turtle exploitation in 44 of the world's 58 marine turtle
RM Us (Figure 3; Dataset S2). The RMUs with the highest exploitation
for each species between 1990 and 2010 included the Southwest
Pacific green turtle RMU, the West Pacific/Southeast Asia hawksbill
RMU, the Northeast Atlantic loggerhead RMU, the East Pacific
olive ridley RMU, the Southeast Atlantic leatherback RMU, and the
Northwest Atlantic Kemp's ridley RMU, the latter of which is the
only RMU for Kemp's ridley turtles (Figure 3; Dataset S2).
In the 1990s, 22 RMUs had documented exploitation, with the
West Pacific/Southeast Asia hawksbill RMU having the highest
number of turtles (n = 750 per year; Figure 3; Dataset S2). In the
2000s, 34 RMUs had documented exploitation, with the Southeast
Indian green turtle RMU having the most exploited turtles (n = 1001
per year; Figure 3; Dataset S2). Finally, the 2010s had documented
exploitation in 28 RMUs, with the most turtles exploited from the
West Pacific/Southeast Asia hawksbill RMU (n = 922 per year;
Figure 3; Dataset S2).
Of the 44 RMUs with documented exploitation, 30 had docu-
mented exploitation in multiple decades, allowing for the analysis of
trends over time. In all, 10 RMUs had documented exploitation in all
three decades, but none showed consistent trends in exploitation
across the entire period (Figure 3; Dataset S2). In total, 20 RMUs
had documented exploitation in only two decades, 7 of which had
exploitation increase over time, 10 of which had exploitation de-
crease, and 3 of which exhibited no change in exploitation (Figure 3;
Dataset S2). Among the RMUs with an overall increase in exploita-
tion, four were “high- risk” populations (West Pacific/Southeast Asia
hawksbill RMU, East Atlantic hawksbill RMU, Southwest Atlantic
hawksbill RMU, South Caribbean green RMU; Dataset S2).
3.4 | RMU exploitation impact scores
Across the entire study period, we were able to calculate exploita-
tion impact scores for 29 RMUs, including two with high exploita-
tion (East Atlantic green and East Atlantic hawksbill RMUs) and two
with moderate exploitation (East Pacific green and South Caribbean
green RMUs; Dataset S2). Over the 30- year study period, the East
Atlantic hawksbill RMU was the only RMU classified as a “high risk”—
“high exploitation” RMU, with an exploitation impact score of 1.16,
meaning, on average, the number of turtles exploited within a given
dec ade was greater than the number of nesting females in the popu-
lation, with 1.16 turtles exploited per nesting female per decade.
In the 1990s, we were able to calculate exploitation impact
scores for 17 RMUs. Only the East Atlantic green turtle RMU
had high exploitation (3.5 turtles exploited per nesting female
per decade) and only the Southwest Pacific hawksbill RMU had
moderate exploitation (0.6 turtles exploited per nesting female
per decade; Figure 4a; Dataset S2). However, both of these RMUs
are considered to be low- risk populations (Wallace et al., 20 11).
In the 2000s, exploitation impact scores were calculated for 23
RMUs, of which four had high exploitation and two had moderate
exploitation (Figure 4b; Dataset S2). Two of the RMUs with high
exploitation in the 2000s— the East Atlantic hawksbill RMU (3.1
turtles exploited per nesting female per decade) and the North
Pacific loggerhead RMU (1.4 turtles exploited per nesting female
per de cade)— were classif ied as “high risk”— “hig h exploit ation.” We
FIGURE 1 Global maps depicting the magnitude of illegal
exploitation by country during the 1990s, 2000s, and 2010s. Data
from the documented literature (i.e., peer-reviewed, grey literature,
and media reports) are shown by colored countries, while data
from the in-country expert online questionnaire are shown with
the diamond symbols (no questionnaire data were collected for the
1990s). Country-specific data represented herein are presented in
Dataset S1. Map lines delineate study areas and do not necessarily
depict accepted national boundaries. [Color figure can be viewed at
wileyonlinelibrary.com]

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SENKO et al.
were able to calculate exploitation impact scores for 19 RMUs in
the 2010s, none of which had high exploitation and only one of
which had moderate exploitation (0.87 turtles exploited per nest-
ing female per decade, the “high risk” South Caribbean green tur-
tle RMU; Figure 4c; Dataset S2).
3.5 | Global distribution and trends in exploitation
by countries
We obtained peer- reviewed publications from 25 countries or ter-
ritories, grey reports from 27 countries or territories, and archived
media reports from 21 countries or territories published between
1990 and 2020, covering 51 countries when combined (herein re-
ferred to as documented sources). The mean data quality score for
these 51 countries ranged from 0.17 to 6.83, with an average of 1.15.
The majority of countries (n = 33, 65%) had a data quality score less
than 1, indicating low data availability and/or quality. Likewise, we
received questionnaire responses corresponding to 34 countries,
including 13 countries for which we found no documented exploi-
tation (Table S1). In total, we acquired estimates of illegal marine
turtle exploitation from at least one data source for 65 countries
and territories. When all data sources are taken into account, nearly
75% of exploitation (i.e., number of turtles exploited) between 1990
and 2020 occurred in five countries: Haiti (31%), Tanzania (20%),
Honduras (10%), Indonesia (7%), and Mexico (6%; Dataset S1). Of
these five countries, Haiti, Honduras, and Tanzania were only repre-
sented in the questionnaire data.
Of the 65 countries with available data, 43 had exploitation es-
timates across multiple decades, allowing for the analysis of trends
over time. In all, 14 countries had estimated increases in exploita-
tion over time, with Mexico having the largest rise in exploitation
(~60,000 turtle increase; Figure 1; Table S3). In total, 15 countries
had estimated decreases in exploitation over time, with Tanzania
having the largest decrease in exploitation (120,000 turtle decrease;
Figure 1; Table S3). Colombia, the Turks and Caicos Islands, and the
United States did not have clear trends in exploitation (Figure 1;
Table S3). The remaining 11 countries had no change in exploitation
estimates (Figure 1; Table S3). Discrepancies between documented
and questionnaire data are presented in Table S4.
3.6 | Global distribution and trends in international
trafficking by countries
Of the 198,297 total marine turtles reported in documented data
between 1990 and 2020, 42,771 (22%) were reported to have been
traded internationally (questionnaire data are not applicable here,
as trafficking was not reported in the questionnaire). Across all
three decades, the top five places of origin included Vietnam (34%),
the Philippines (8%), Venezuela (4%), Malaysia (3%), and the Coral
Triangle (2%), the latter of which could include Indonesia, Malaysia,
the Philippines, Papua New Guinea, Timor- Leste, and the Solomon
Islands. The top five destinations for traded marine turtles included
China (46%), Japan (43%), Colombia (4%), Vietnam (2%), and Malaysia
(2%). The five most common routes for international trade between
1990 and 2020 were Vietnam to China (34%), the Philippines to
China (5%), Venezuela to Colombia (4%), the Coral Triangle to China
(2%), and Malaysia to China (2%).
There was no clear decadal trend regarding the number of turtles
trafficked, with 18,022 turtles traded in the 1990s, 4776 traded in
the 2000s, and 19,973 traded in the 2010s. Four countries were doc-
umented as origins of marine turtle trafficking in multiple decades,
with the number of trafficked turtles originating from Indonesia and
Malaysia decreasing over time and trafficked turtles coming from
the Philippines and Vietnam increasing. Six destination countries
had documented trafficking in multiple decades, with Australia,
Indonesia, and Japan showing a decrease in imported turtles and
China, Vietnam, and Russia showing an increase in imported turtles.
Five trafficking routes showed trends in the number of turtles traf-
ficked over time. Indonesia to China and Malaysia to China showed
reduced trafficking of marine turtles, while the Philippines to China,
the Philippines to Vietnam, and Vietnam to China all showed in-
creased trafficking over time.
FIGURE 2 Total number of marine turtles illegally exploited annually by species in the (a) 1990s, (b) 2000s, and (c) 2010s based on an
assessment of documented sources (online questionnaire data were not included due to the absence of species- specific data). Species on
the x- axis are abbreviated as green turtle (Cm), hawksbill turtle (Ei), loggerhead turtle (Cc), olive ridley turtle (Lo), and leatherback turtle (Dc),
while turtles not reported by species are denoted as NR. Species- specific data represented herein are presented in Dataset S1. [Color figure
can be viewed at wileyonlinelibrary.com]

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4 | DISCUSSION
4.1 | Overview and caveats of illegal marine turtle
exploitation
Our assessment revealed that over 1.1 million marine turtles were
reported to be illegally exploited in 65 countries and territories and
in 44 of the world's 58 marine turtle RMUs between 1990 and 2020,
with over 44,000 turtles exploited per year during the past dec-
ade. While there was no consistent trend in the number of turtles
exploited across all three decades, there was a 28% decrease in
reported exploitation from the 2000s to the 2010s. Green turtles
(56%) and hawksbills (39%) comprised most of the reported exploi-
tation across all three decades. With a few exceptions highlighted
herein, most of the reported exploitation over the past decade oc-
curred in ‘low- risk’ marine turtle RMUs that are typically genetically
diverse and characterized by large, relatively stable, or increasing
abundances, suggesting that current levels of illegal exploitation— at
least relative to other contemporary threats— may not exert major
population- level impacts on most marine turtle RMUs (Figure 4).
FIGURE 3 Global maps for green, hawksbill, olive ridley, loggerhead, and leatherback turtle Regional Management Units (RMUs)
showing the total number of recorded turtles illegally exploited for each species' respective RMUs across the three decades based on an
assessment of peer- reviewed literature, grey literature, and media reports (online questionnaire data were not included due to the absence
of species- specific data). RMUs are defined as “georeferenced available data on marine turtle biogeography— including individual nesting
sites, genetic stocks, and geographic distributions based on monitoring research— to develop multi- scale RMUs that spatially integrate
sufficient information to account for complexities in marine turtle population structures” (Wallace, DiMatteo, et al., 2010, see for a full list
of the 58 global marine turtle RMUs). RMU data represented herein are presented in Dataset S2. Map lines delineate study areas and do not
necessarily depict accepted national boundaries. [Color figure can be viewed at wileyonlinelibrary.com]

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SENKO et al.
While we may have missed obscure literature or media re-
ports and were unable to reach all of the experts we sought, we
are confident that our assessment generally reflects the current
state of available information regarding illegal exploitation of
marine turtles worldwide. Nonetheless, our estimates of marine
turtl e exploi tat ion are almost certa inly biased nega tivel y for sev-
eral reasons, including (1) assessing prohibited exploitation is dif-
ficult due to its typically clandestine nature (Senko et al., 2 014;
von Essen et al., 2 014); (2) there are significant data gaps due to
a lack of reporting in many areas, limited spatial scope of assess-
ment, or difficulty accessing obscure data (e.g., programmatic
grant or contract reports) that were not searchable or available
for our assessment; (3) data represented herein are likely more
indicative of hotspots where researchers have published data,
as well as extemporaneous repor ts of trade seizures, than actual
illegal exploitation; (4) international trade of marine turtles may
be on the rise due to increasing demand from Asia, and espe-
cially China (Yifan, 2018); (5) turtles may be butchered at sea in
nearshore waters for meat or products, and stranding probabil-
ities of marine turtle carcasses tend to be low (usually 5%– 30%
of total mortality; Epperly et al., 1996; Hart et al., 20 06; Koch
et al., 2013), limiting detection of exploitation; (6) a general lack
of law enforcement leads to limited successful documentation
of illegal exploitation (Mancini et al., 2011); (7) our online ques-
tionn air e and use of me dia so urc es wer e cond ucte d and ass essed
in English, representing a potential language bias that likely re-
sulted in missed coverage from non- English- speaking countries;
and (8) our assessment included only whole turtles or products
that could be linked to individual turtles (e.g., heads, shells, or
tails), and did not account for eggs or products that could not be
linked to individual turtles (e.g., bracelets or earrings made from
tortoiseshell).
4.2 | Trends in illegal marine turtle exploitation
over the past three decades
Although there was no consistent trend in the number of turtles
exploited across all three decades, there was a 28% decrease in re-
ported exploitation from the 2000s to the 2010s. Considering that
more information was available for the most recent decade com-
pared to the 2000s (we found more sources documenting illegal
marine turtle exploitation and trade for the 2010s than the 1990s
and 2000s combined; Table S2), we expected an overall increase in
the magnitude of reported exploitation. However, we observed the
opposite, suggesting that the reported decrease in exploitation over
the past decade may indicate an actual decrease. This could be due
to a combination of increased protective legislation, awareness of
the problem, global grassroots conservation efforts, changing local
norms and traditions, reduced research attention or shifting con-
servation priorities (especially for emerging threats such as plastic
pollution), fewer documented exploitation interventions or quanti-
fications, or declining marine turtle populations (i.e., fewer turtles
available for exploitation) at local scales. Decreased exploitation
may also be attributed to greater avoidance of the activity due to
increased protective legislation or risks to assessing the activity,
especially if it becomes more organized or connected with crime
syndicates.
There were inconsistent trends concerning the species ex-
ploited, RMUs, countries, and marine turtle trafficking. While ex-
ploitation across the entire study period consistently comprised
mostly hawksbill and green turtles, exploitation in the 1990s was
primarily hawksbills, while the 2000s experienced a large spike in
green turtle exploitation (Figure 2). This may be due to lingering
effects of shifting legislation in bekko trade, increases in green
turtle populations (Broderick et al., 2006), and discrepancies in
FIGURE 4 Risk and exploitation impact scores for each Regional Management Unit (RMU) with documented exploitation in the (a) 1990s,
(b) 2000s, and (c) 2010s. RMUs are defined as “georeferenced available data on marine turtle biogeography— including individual nesting
sites, genetic stocks, and geographic distributions based on monitoring research— to develop multi- scale RMUs that spatially integrate
sufficient information to account for complexities in marine turtle population structures” (Wallace, DiMatteo, et al., 2010, see for a full list
of the 58 global marine turtle RMUs). The RMU in each decade with the highest exploitation impact score is highlighted. To calculate the
exploitation impact scores, estimates for the total number of turtles exploited from each RMU in each decade were divided by estimates
of nesting females from the respective RMU. Risk scores and estimates of nesting females were sourced from Wallace et al. (2011). RMU
conservation priorities increase moving toward the upper- right corner of the plot. RMU exploitation impact scores represented herein are
presented in Dataset S2. [Color figure can be viewed at wileyonlinelibrary.com]

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data collection (e.g., the 2000s having a much higher proportion
of exploitation reported by species compared to the other two
decades).
Trends in the number of marine turtles exploited in each coun-
try were also inconsistent, with Mexico having the largest in-
crease in exploitation and Tanzania having the largest decrease
in exploitation. Country trends were the only trends, besides
overall exploitation, that could be informed by both documented
and questionnaire data. While this provides a more complete pic-
ture of exploitation, using all data sources to assess trends in a
country's exploitation has some important caveats. For example,
questionnaire data reported consistently higher exploitation than
documented data (Table S4), which may lead to incorrect trends
for countries that utilize both documented and questionnaire data
for trend analysis. However, this issue is only per tinent to Guinea-
Bissau, Jamaica, and Mexico, all of which showed increased ex-
ploitation in the ana lysis with both documented and questionnaire
data, but decreased exploitation from questionnaire data alone
(Table S4).
While there was no clear decadal trend in the number of tur-
tles trafficked internationally from 1990 to 2020, trend analysis
revealed that patterns of trafficking shifted in Indonesia and
Vietnam. Over the course of three decades, the number of tur-
tles exported from Indonesia to other countries decreased, but
the number of turtles imported to Indonesia increased during
the same period. Vietnam has shifted in the opposite direction,
with more turtles being exported and fewer imported in recent
times co mpared to the 1990s . Ho wever, traf f i cking tr ends sh ould
be interpreted with caution for several reasons. For example,
it was not always known whether marine turtles were traded
domestically or internationally, nor whether seizures were a re-
sult of international trade or local use, an issue that could be
resolved using forensic testing to determine a marine turtle's
origin. Furthermore, many instances of international trade are
not documented, particularly in cases involving transport of
marine turtle products on the high seas or open oceans, where
avoiding detection (akin to illegal, unregulated, and unreported
fisheries) is easier than in cases where turtles are exploited in
coastal areas.
4.3 | Conservation priorities based on RMU
exploitation impact scores
Assessment of exploitation by RMUs based on documented data
and exploitation impact scores allows for the identification of
“high risk” and “high exploitation” global hotspots, which can in-
form conservation and management priorities (Barrios- Garrido
et al., 2020; Wallace, DiMatteo, et al., 2010; Wallace et al., 2011,
2013). At the RMU scale, most of the RMUs with increasing ex-
ploitation are “low- risk” populations, while some “high- risk” RMUs
(i.e., hawksbill RMUs in the West Pacific/Southeast Asia, East
Atlantic, and Southwest Atlantic, South Caribbean green turtle
RMU) also experienced increases in documented exploitation. The
only RMU that emerged with a “high risk” and “high exploitation”
conservation priority across the three- decade period of this study
was the East Atlantic hawksbill RMU. Thus, our analysis showed
that most of the world's exploitation over the past decade has
been documented in “low- risk ” marine turtle RMUs th at tend to be
genetically diverse and exhibit large, stable, or increasing popu-
lation abundances. Our analysis, therefore, suggests that illegal
exploitation may not represent a major, population- level threat to
most of the world's marine turtle RMUs, with some exceptions.
These exceptions include the four “high- risk” RMUs that exhib-
ited increasing trends in exploitation over time (i.e., West Pacific/
Southeast Asia hawksbill RMU, East Atlantic hawksbill RMU,
Southwest Atlantic hawksbill RMU, and South Caribbean green
RMU). However, we note that our assessment is based on limited
data availability and estimates are likely negatively biased and
geographically incomplete.
Despite the utility of RMUs and exploitation impact scores for
assessing conservation priorities for marine turtle populations, these
results should be interpreted cautiously for the following reasons.
First, RMUs could only be assigned to turtles that were reported
by species, which excluded online questionnaire data and 34% of
documented data. Second, the location of capture is also necessary
to assign an RMU. Thus, for trafficked turtles seized en route to or
at their destination, an RMU could not be assigned unless the lo-
cation of capture was specifically stated. Furthermore, some coun-
tries such as Malaysia, Indonesia, and the Philippines are included
within the boundaries of multiple RMUs, making it difficult to assign
the correct RMU unless the exact location of capture or the genetic
stock to which turtles belong to are known. Third, the exploitation
of marine turtles at higher risk RMUs may be more difficult to detect
due to low population numbers and potentially stricter enforcement
or penalties. Additionally, the RMU classifications and population
estimates used for the analysis were published a decade ago (i.e.,
Wallace et al., 2011), though drastic changes in nesting abundance
within one decade that would substantially alter the conclusions of
our study were unlikely because marine turtles are long- lived and
their population sizes vary across large temporal and spatial scales.
Another limitation of sourcing data from Wallace et al. (2011) is that
exploitation impact scores could not be calculated for some RMUs
because no estimate of female nesting abundance was available.
With these limitations in mind, RMU- scale evaluations of poten-
tial population- level impacts of exploitation enabled identification
of conservation priorities as well as knowledge gaps that require
attention.
4.4 | Marine turtle trafficking and hawksbill turtle
exploitation as a global conservation priority
When our analysis is combined with the most recent global assess-
ment of legal exploitation that reported an annual take of greater
than 42,000 marine turtles since 2010 (Humber et al., 2014), at least

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SENKO et al.
approximately 80,000 turtles are exploited— legally or illegally—
per year, most of which are green and hawksbill turtles. Humber
et al. (2014) found that >95% of legal marine turtle exploitation was
comprised of green (89%) and hawksbill turtles (8%). However, in
the present analysis, hawksbills made up over half of the illegal ex-
ploitation where species was reported over the past decade— more
than six times higher than the proportion of hawksbills reported in
Humber et al. (2014). Regardless, our estimate underestimates the
true exploitation of hawksbills given that we were unable to ascribe
tortoiseshell products (e.g., pieces of jewelry) to individual turtles.
A recent global assessment of the tortoiseshell trade reported that
more than 46,0 00 individual tortoiseshell products have been of-
fered for sale since 2017 (>17,000 in- person and >29,000 online)
in at least 10 countries with substantial illegal markets (Nahill
et al., 2020).
Madagascar and countries in Asia (i.e., China, Vietnam,
Indonesia, Malaysia, and the Philippines) emerged as contem-
porary hawksbill exploitation hotspots in our study. Based on
known cases that resulted in hundreds of animals confiscated
from vessels over the past decade, it is believed that vessels
leave ports with the express purpose of targeting hawksbills from
Malaysia, the Philippines, and Indonesia to supply international
black market circuits, including high- end consumer markets of
East Asia, with theoretical flight and vessel routes flowing from
these countries directly to China or through Vietnam into China
(CITES Secretariat, 2019; Gomez & Krishnasamy, 2019). Anecdotal
evidence suggests that hawksbill turtles in Asia are targeted by
vessels originating from the Hainan Province in China as well as
Vietnam and Thailand (CITES Secretariat, 2019; Lam et al., 2012),
while Africa may be supplying marine turtles seized in China via
transshipment through Indonesia (anonymous survey respondent;
this study).
Traded turtles may also be obtained directly from fishers who
capture turtles as bycatch and then warehouse them to supply black
market circuits (Yifan, 2018). Tortoiseshell is nonperishable and thus
easily stockpiled, which distances consumers from suppliers, re-
duces traceability, and decouples the feedback of reduced demand
with increased cost over value (McClenachan et al., 2016; Miller
et al., 2019; Nahill et al., 2020). For example, in 2014, Vietnamese
authorities seized 7000 turtle carcasses in a warehouse, the single
largest seizure of marine turtles ever recorded; hawksbill turtles
comprised the majority of carcasses, almost all of which were either
fully or partially taxidermied and assumed to be bound for sale in
China (Nuwer, 2016).
For these reasons, special attention should be paid to hawks-
bill RMUs, particularly the West Pacific hawksbill (Pacific Ocean/
Southeast Asia) RMU. Despite the lack of an exploitation impact
score (due to lack of population abundance data), this population
should be considered another high- priority RMU due to the large
number of turtles exploited, the region's high concentrat io n of cou n-
tries that serve as sources and sinks for international trade, the large
number of individual tortoiseshell products documented in a recent
global assessment (Nahill et al., 2020), and the global conservation
status of hawksbills (Barrios- Garrido et al., 2020; Mortimer &
Donnelly, 2008; Spotila, 20 04).
4.5 | Contextualizing illegal marine turtle
exploitation with other threats
Contextualizing exploitation with other marine turtle threats is
vital for strategic conservation priority- setting. Wallace, Lewison,
et al. (2010) reported that published information documented
a minimum of 85,000 turtles reported as bycatch globally from
1990 to 2008, a value that the authors suggested was underesti-
mated by at least two orders of magnitude due to low (<1%) ob-
served international fishing effort and underrepresented bycatch
in small- scale fisheries. In the Mediterranean alone, Casale (2011)
estimated that 44,000 turtles are killed annually as bycatch. In
contrast, illegal exploitation may not be comparable to bycatch
impacts on a global scale, although better data for both bycatch
and exploitation are needed. Regardless, most populations im-
pacted by exploitation appear to be from RMUs characterized by
relatively high population sizes and stable or increasing population
trends, with the likely exceptions of hawksbill turtles in the West
Pacific/Southeast Asia and Caribbean. Nonetheless, because ex-
ploitation often affects older life stages (i.e., large juveniles and
adults) that have the largest influence on marine turtle population
dynamics (Crouse et al., 1987), it can be a serious threat to marine
turtle populations.
4.6 | Addressing illegal marine turtle
exploitation and future research considerations
To successfully address illegal marine turtle exploitation, a combina-
tion of outreach, enforcement, and research is necessary, particu-
larly to protect “high- risk” marine turtle populations from further
declines. However, before such strategies can be developed, how
“illegal” hunting is defined, described, and quantified must be con-
sidered carefully. Specifically, enhanced culturally appropriate out-
reach is crucial to balancing goals of sustaining turtle populations
and preser ving important cultural and socioeconomic traditions. For
example, indigenous Australians have a legally recognized right to
use marine turtles for food and other customary activities within
their ancestral territories (Weiss et al., 2013); similar characteriza-
tions likely exist elsewhere. Furthermore, turtle exploitation can
help sustain food insecure or otherwise marginalized communities,
and prohibiting exploitation or fully enforcing laws that prohibit
exploitation will likely be detrimental to human welfare (Barrios-
Garrido et al., 2019; Liles et al., 2015).
Moving forward, increased support for governments lacking
the resources to properly enforce illegal exploitation of marine
turtles is needed, as well as support for communities to sustain
human well- being in the face of restrictions or bans on marine
turtle exploitation. Communities living near marine turtle supply

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S ENKO e t al.
centers could be provided with incentives— that is, investments in
he a lthcar e , educat ion , and infr ast ruc tur e , as well as in c reas ed ec o -
nomic opportunities— to protect wildlife while sustaining import-
ant cultural practices (Challender & MacMillan, 2014; Gjertsen &
Niesten, 2010; Pakiding et al., 2020). Co- design of legislation be-
tween stakeholders and governments could offer a pathway for
developing conservation strategies that benefit both people and
turtles.
Further data collection and reporting by governments and
enforcement agencies will be essential for researchers and man-
agers to accurately prioritize conservation efforts to address il-
legal exploitation in the context of other threats to marine turtle
populations, particularly in countries that are lacking robust data
regarding exploitation, which comprised the majority of coun-
tries in our assessment (Dataset S1). Improving assessments of
exploitation (and other threats) are needed to better understand
population- level impacts to marine turtles. Future research should
strive to develop population models that estimate the number of
turtles of harvestable size in each RMU, while enforcement offi-
cers could collect genetic data on seized marine turtles to allow
for correct RMU attribution or record size classes to better under-
stand population- level effects of exploitation. Further research
looking at how marine turtle exploitation relates to factors such
as population size, food insecurity, and the Human Development
Index will help researchers elucidate the drivers behind supply and
demand for marine turtle products (Barrios- Garrido et al., 2020).
Additionally, given the importance of public perceptions, greater
conservation and regulatory buy- in by local government and influ-
ential community leaders could be beneficial (Liles et al., 2016; von
Essen et al., 2 014). Traditional practices (i.e., taboo systems or cul-
tural rules) which once encouraged the protection of and respon-
sibility to care for sensitive or depleted resources are currently
experiencing a renaissance in numerous Pacific Island locations
(Summers et al., 2018), and may be replicated elsewhere as well as
complement current management efforts.
With effective conservation and management, marine turtle pop-
ulations have shown remarkable potential for recovery from long-
term abundance declines caused by persistent threats (Chaloupka
et al., 2008; Mazaris et al., 2017). While the results herein almost
certainly represent only a fraction of actual illegal exploitation, this
study provides a foundation for further research and conservation
efforts focused on addressing the illegal exploitation of marine tur-
tles throughout the world's oceans.
ACKNOWLEDGMENTS
We thank the following individuals/organizations (who wished to
be acknowledged) for participating in our online questionnaire:
Manish Chandi, Linda Searle, Maria Angela Marcovaldi, Eduardo
HSM Lima, Fundação Pró- TAMAR, Janice Blumenthal, Karla G.
Barrientos, Cristian Ramirez- Gallego, Michael White, Boomerang for
Earth Conservation, Fiji Department of Fisheries, Paulo Catry, Jean
Wiener, Stephen G. Dunbar, Kartik Shanker, Udayana University
(Bali, Indonesia), White River Watershed, Ocho Rios Marine Park
Association, Blue Ventures Conservation, Nicolas Pilcher, Thomas
Le Berre, Vicente Guzmán Hernández, Grupo Tortuguero de
las Californias, Tyffen Read, Jeff Kinch, ProDelphinus, Teodora
Bagarinao, Portugal Centre of Marine Species, Jessica Berkel,
Claire Saladin, Lindsey West, Thomas Stringell, Claudia Lombard,
Liza Boura, and Jillian Hudgins. We thank Karen Eckert for help
distributing the online questionnaire to The Wider Caribbean Sea
Turtle Conservation Network (WIDECAST) country experts and
the School for the Future of Innovation in Society at Arizona State
University for providing Kayla Burgher with an Undergraduate
Research Fellowship. Wallace J. Nichols inspired the idea behind
this paper. Casey Lyn, Sydney Hitzig, Suzanne Higgs, David Shockey,
Alexis Janke, and Janie Reavis assisted with data collection. John
Wang, Michelle María Early Capistrán, Jose Urteaga, Hector Barrios,
Annette Broderick, and three anonymous reviewers provided help-
ful comments and suggestions that substantially improved the final
manuscript. The findings and conclusions in this manuscript are
those of the authors and do not necessarily represent the views, of-
ficial policy, or position of any agency of the U.S. Government.
CONFLICT OF INTEREST
The authors declare no competing interests.
DATA AVA ILAB ILITY STATE MEN T
The data that support the findings of this research are available in
the Open Science Framework at https://osf.io/n8uhr/
ORCID
Jesse F. Senko https://orcid.org/0000-0002-3228-4218
Kayla M. Burgher https://orcid.org/0000-0002-5291-1039
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SUPPORTING INFORMATION
Additional supporting information can be found online in the
Supporting Information section at the end of this article.
How to cite this article: Senko, J. F., Burgher, K. M., del Mar
Mancha- Cisneros, M., Godley, B. J., Kinan- Kelly, I., Fox, T.,
Humber, F., Koch, V., Smith, A. T., & Wallace, B. P. (2022).
Global patterns of illegal marine turtle exploitation. Global
Change Biology, 28, 6509–6523. https://doi.org/10.1111/
gcb.16 378