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Freshwater eels: A symbol of the eﬀects of global change
Fish and Fisheries · July 2018
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1 | INTRODUCTION
1.1 | Global change: Five components that threaten
In 2005, the Millennium Ecosystem Assessment (2005a) pointed
out that, despite an increase in human welfare during the twentieth
century, anthropogenic actions threaten the ability of ecosystems
to sustainably provide important goods and services, especially for
future generations. Such human- caused environmental changes are
generally referred to as global change (Steffen et al., 2005). In this
paper, we illustrate how the rate at which global change is happen-
ing can endanger even highly adaptive species such as temperate
eels Anguilla anguilla (European eel), A. rostrata (American eel) and
A. japonica (Japanese eel). In a more specific manner, this example
illustrates that cumulative effects of the different components of
global change (Jacoby et al., 2015; Miller, Feunteun, & Tsukamoto,
2016; Tylianakis, Didham, Bascompte, & Wardle, 2008; Vitousek,
D’antonio, Loope, Rejmanek, & Westbrooks, 1997; Western, 2001)
can produce rates of change that exceed a species’ adaptive capac-
ity, a central question in the debate on the effects of climate and
global change (Donner, Skirving, Little, Oppenheimer, & Hoegh-
Guldberg, 2005; Visser, 2008).
Soulé (1991) proposed a list of the main threat s to biodiversity
that include six components: habitat loss, habitat fragmentation,
overexploitation, exotic species, pollution and climate change. Later,
the United States National Research Council (200 0) proposed a very
similar list of ongoing changes: (a) climate, (b) land use and land cover
modifications that can result in habitat loss and fragmentation, (c)
biogeochemical and hydrological cycles and pollution, (d) biotic mix-
ing including biological invasions, and (e) overexploitation of natu-
ral resources especially in oceanic ecosystems. The main difference
with Soulé’s proposal is that habit at loss and habitat fragmentation
were merged into a single component. This lat ter listing was then
DOI : 10.1111 /fa f.123 00
Freshwater eels: A symbol of the effects of global change
Hilaire Drouineau1 | Caroline Durif2 | Martin Castonguay3 | Maria Mateo1 |
Eric Rochard1 | Guy Verreault4 | Kazuki Yokouchi5 | Patrick Lambert1
1Irstea, UR EABX, centre de Bordeaux,
Cestas Cedex, France
2Instit ute of Mari ne Resea rch, Stor ebø,
3Ministère des Pêches et des
Océans, Institut Maurice-Lamontagne,
Mont-Joli, QC, Canada
4Ministè re des Forêts, de l a Faune et des
Parcs – Direction Régionale du Bas Saint-
Laurent, Rivière-du-Loup, QC, Canada
5Nationa l Resear ch Institute of Fis heries
Science , Fisher ies Rese arch Agency,
Kanazawa, Yokohama, Kanagawa, Japan
Hilaire Drouineau, Irstea, UR EABX, centre
de Borde aux, 50 avenue de Verdun, F-33612
Cestas Cedex, France.
Japan Fisheries Research and Education
Temperate eels Anguilla anguilla (European eel), A. rostrata (American eel) and A. ja-
ponica (Japanese eel) are three catadromous species which have been declining since
the 1970s/1980s despite their remarkable adaptive capacity. Because of their spe-
cific life cycles, which share distant oceanic spawning grounds and continental
growth stage, eels are affected by five components of the global change: (a) climate
change affecting larval survival and drift, (b) an increase in pollution leading to high
levels of contamination exacerbated by their high lipid levels, (c) increasing fragmen-
tation and habitat loss that reduce dramatically the amount of available habitats and
induce increased spawner mortality, (d) the appearance of Anguillicola crassus a para-
sitic alien nematode that impairs spawning success, and (e) the impact of commercial
and recreational fisheries for all life stages of eel. In this context, the rapid increases
of pressures during the “Great Acceleration” have surpassed the adaptive capacity of
eels. This illustrates that cumulative effects of global change can lead to the collapse
of species, even in species that have amazingly high adaptive capacities.
adaptation, Anguilla spp., climate change, contamination, ecosystem fragmentation,
DROUINEAU Et Al .
endorsed by many authors, including IPCC (2001), the Millennium
Ecosystem Assessment (2005b), Simberloff (2012) and Pe’er et al.
(2013), although Pe’er et al. (2013) did not mention pollution. The
IGPB (Steffen et al., 2005; executive summary) and the Global
Change Program of the Royal Societ y of Canada (1992) provided
more detailed lists of components of global change (IGBP: oil har-
vest, transformation of land surface, nitrogen waste, use of fresh-
water, greenhouse gas, marine habitat destruction, overexploitation
of fisheries, extinction rates of species—GCPRSC: climate change,
energy and resource consumption, air and water pollution, ozone de-
pletion, population increase, extinction events, land and soil degra-
dation). These components can be merged into the five more general
components from IPCC (except perhaps “population increase” which
can be shared by various components).
Among these components of global change, the best documented
is global warming, which is due to increased greenhouse gas emission
by human activities (IPCC, 2015), affecting the biosphere at all scales
(Gattuso et al., 2015). Global warming already has visible impacts on
the ecology of living organisms (Hughes, 200 0; Walther et al., 2002)
with impacts on their phenology (Chevillot et al., 2017; Menzel et al.,
2006) and modification of distribution areas (Cheung et al., 2010;
Lassalle, Crouzet, & Rochard, 2009; Nicolas et al., 2011; Rougier
et al., 2015). Global warming alters biogeochemical cycles such as the
water cycle. For example, the seasonalit y of river discharge (ampli-
tude between low and high discharge) is expected to increase, while
average discharges will increase in some regions and decrease in oth-
ers (Nohara, Kitoh, Hosaka, & Oki, 2006; van Vliet et al., 2013).
The second component is increasing nutrient, contaminant and
pesticide loads in the ecosystem due to industries, agriculture and
urbanization (Verhoeven, Arheimer, Yin, & Hefting, 2006). These
increased loads destabilize nutrient cycles and can have direct con-
seq uence s,s uchase utrop hicat ion(Raba lais,Turner,Dí az,&Ju st ić,
2009; Tilman et al., 2001). At the individual scale, contaminants
and pesticides can have a large range of deleterious ef fect s, such
as altered metabolism, immunotoxicity, endocrine disruption or
neurotoxicity (ICES, 2016; Köhler & Triebskorn, 2013). Moreover,
many contaminants are ecologically harmful because persistent
biomagnifying chemicals can accumulate in food webs at high tro-
phic levels (Köhler & Triebskorn, 2013; Van Oostdam et al., 2005).
Another component is the modification of habitats due to anthro-
pogenic land use, which can lead to fragmentation of aquatic and ter-
restrial ecosystems or even habitat loss (Brook, Sodhi, & Bradshaw,
2008; Collinge, 1996; Fischer & Lindenmayer, 2007). Habitat loss and
ecosystem fragmentation are currently considered major threats to
biodiversity and represent one of the major challenges in ecosystem
conservation and restoration (Sutherland et al., 2013; Tilman, May,
Lehman, & Nowak, 1994; Tischendorf & Fahrig, 20 00a, 200 0b), by
impairing the ability of individuals to migrate to essential habitats
(Gros & Prouzet, 2014), by isolating populations and reducing gene
flow (Haxton & Cano, 2016; Horreo et al., 2011), and by modifying
species community structure (Perkin & Gido, 2012). Fragmentation
and habitat loss increase the risk of extinction cascades (Fischer &
Lindenmayer, 2007; Haddad et al., 2015; Junge, Museth, Hindar,
Kraabøl, & Vøllestad, 2014; Krauss et al., 2010; van Leeuwen,
Museth, Sandlund, Qvenild, & Vøllestad, 2016; Terborgh et al., 20 01).
Biological invasion by alien species is our fourth component of
global change (Occhipinti- Ambrogi & Savini, 2003; Ricciardi, 2007;
Vitouse k et al., 1997). Invasive s pecies can affe ct native ones dire ctly
through predation, competition or parasitism, or indirectly by habi-
tat modification or by spreading diseases (Lymber y, Morine, Kanani,
Beatt y, & Morgan, 2014). Arrival of alien species due to shif ts in their
distribution in response to climate change, uniformization of habitat
or transport by humans can profoundly reshape species interactions
and has consequences at the species, community and ecosys tem lev-
els but also on the provision of ecosystem ser vices (Alpine & Cloern,
1992; Cloern, 1996; Grosholz, 20 02; Vilà et al., 2010). Now, there
are about 10,000 alien species registered in Europe, while ecological
impact s have been documented for 11% of them and economic im-
pacts for 13% (Vilà et al., 2010).
1 INTRODUCTION 1
1.1 Global change: Five components that threaten
1.2 Temperate eels: Endangered species impacted
by diverse anthropogenic pressures
2 COMPONENT 1—GLOBAL WARMING AND
OCEAN MODIFIC ATION: IMPACTS ON EEL
2.1 Effect on leptocephalus drift and survival 5
2.2 Effect on silver eel spawning migration 7
3 COMPONENT 2—INCREASED
CONTAMINATION LOAD: CONTAMINATION OF
EELS AND CONSEQUENCES ON
4 COMPONENT 3—FRAGMENTATION AND
HABITAT LOSS: FRAGMENTATION BY WEIRS
AND DAMS AND CONSEQUENCES ON
UPSTREAM AND DOWNSTREAM MIGRATION
4.1 Movements, habitats and fragmentation 8
4.2 Blockage during upstream migrations 9
4.3 Impaired downstream migrations 9
5 COMPONENT 4—ALIEN SPECIES: EFFECTS OF
ALIEN PARASITOID Anguillicola crassus
6 COMPONENT 5—EXPLOITATION OF NATURAL
RESOURCES: AN INTENSIVE E XPLOITATION OF
EELS AT ALL THEIR S TAGES
7 WHEN FORT Y YEARS OF GLOBAL CHANGE
HAS HAD A G REATER IMPACT THAN THE ICE
AGES OR CONTINENTAL DRIFT
8 THE RESILIENCE OF EELS SEVERELY IMPAIRED
BY GLOBAL CHANGE
9 OTHER IMPLICATIONS FOR EEL
MANAGEMENT AND RESEARCH
DROUINE AU Et Al.
At last, the last component is the overexploitation of natural re-
sources (Brook et al., 2008). For example, fisheries have had major
impact s on marine ecosystems, depleting stocks with potential
impact s on fishing food webs, and impacts on habitat s because of
destructive fishing gear (Branch, 2015; Christensen et al., 2003;
Drouineau, Lobr y, et al., 2016; Gascuel et al., 2011; Turner, Thrush,
Hewitt, Cummings, & Funnell, 1999). Fisheries have also shaped
the life histories of exploited fishes by acting as a permanent and
continuous selection pressure (Heino, Pauli, & Dieckmann, 2015;
Jørgensen & Renöfält, 2013).
Through the combined impacts of these five components, global
change leads to an extremely rapid modification of ecosystems.
Species can display different kinds of adaptive response to address
this threat, such as local adaptation through (a) a micro- evolutionar y
response, (b) phenot ypic plasticit y (Charmantier et al., 2008) or (c)
modification of their distribution area (Hughes, 2000). However,
adaptation is not always possible, especially in cases of synergies
among components which can lead to rates of change outpacing a
species adaptive capacity (Brook et al., 2008). In this context, global
change threatens extinction for many species (Brook et al., 2008;
Cahill et al., 2013; Spurgeon, 20 00; Steffen et al., 2005; Thomas
et al., 2004; Urban, 2015). The collapse of the three temperate an-
guillid ee l populations is an excellent illus tration of this phenomenon.
1.2 | Temperate eels: Endangered species impacted
by diverse anthropogenic pressures
European eel, American eel and Japanese eel are three temperate
catad romous species th at share many remark able ecological features
(Figure 1). They have a large distribution area in continental waters
FIGURE1 Life cycle of the three
Anguilla species and effec ts of global
FIGURE2 Spawning grounds (Miller
et al., 2015; Tsukamoto et al., 2011) (open
circles) and continent al distribution of
yellow eels (filled shapes) (Jacoby et al.,
2015) for American (red), European (blue)
and Japanese (green) eels
DROUINEAU Et Al .
(Figure 2; Tesch, 2003), from Canada to Venezuela for American eel
(Helfman, Facey, Stanton Hales, & Bozeman, 1987), from Nor thern
Philippines to Korea for Japanese eel (Tsukamoto, which are reached
after a long larval drift (larvae are called leptocephali) from distant
marine spawning grounds (McCleave, 1993; Schmidt, 1923) and
west of Mariana Islands for Japanese eel (Tsukamoto, 1992)). After
this migration, the larvae metamorphose into glass eels upon reach-
ing the continental shelf (Tesch, 2003). They subsequently penetrate
continental waters, turning into pigmented yellow eels, where they
colonize a large range of continental habitats from brackish to fresh-
water (Arai & Chino, 2012; Daverat et al., 2006). After a growth
phase las ting from 3 to over 30 years, yellow ee ls metamorpho se into
silver eels and migrate back to their spawning grounds (Béguer- Pon,
Castonguay, Shan, Benchetrit, & Dodson, 2015; Chang, Miyazawa,
& Béguer- Pon, 2016; Righton et al., 2016). The eels mature en route
and presumably die following spawning.
Eels have successfully adapted to ver y heterogeneous growth
environments at both the distribution area and catchment scale.
Eels are panmictic (Als et al., 2011; Côté et al., 2013; Han, Hung,
Liao, & Tzeng, 2010; Pujolar, 2013), and their long larval drift limits
the possibility of local genetic adapt ation. However, there are cor-
relations between environmental gradients and spatial patterns in
life history traits throughout the distribution area and within river
catchments (Drouineau, Rigaud, Daverat, & Lamber t, 2014; Vélez-
Espino & Koops, 2009; Yokouchi et al., 2014). For example, sex ratio
is generally female- biased in the northern parts of distribution areas
(Helfman, Bozeman, & Brothers, 1984; Vladykov, 1966; Vøllestad,
1992; Vøllestad & Jonsson, 1988) and in upstream par ts of river
catchments (Oliveira & McCleave, 2000; Tesch, 20 03). Length- at-
silvering (onset of sexual maturation) varies between sexes and
habitats. Males follow a time- minimizing strategy, leaving continen-
tal waters as soon as they reach the minimal length to achieve the
spawning migration, while females follow a size- maximizing strategy,
adapting their length- at- silvering to local growth and mortality con-
ditions finding a trade- off between survival and fecundit y (Helfman
et al., 1987; Vøllestad, 1992). Therefore, male length- at- silvering is
rather set (Oliveira, 1999; Vøllestad, 1992), while females exhibit a
wider range of sizes and are often larger in the northern parts of
the distribution areas (Davey & Jellyman, 2005; Helfman et al., 1987;
Jessop, 2010). These pat terns of life history trait s are thought to be
the result of adaptive phenotypic plasticity that allows individuals
to adapt their life history traits to a wide range of environmental
conditions (Côté, Castonguay, McWilliam, Gordon, & Bernatchez,
2014; Drouineau et al., 2014; Mateo, Lambert, Tétard, Castonguay,
et al., 2017) but also of genetic polymorphism leading to spatially
varying selection and/or genetically based habitat selection pro-
ducing genetically distinct ecotypes (Côté et al., 2014; Gagnaire,
Normandeau, Côté, Hansen, & Bernatchez, 2012; Mateo, Lambert,
Tétard, Castonguay, et al., 2017; Pavey et al., 2015; Ulrik et al., 2014).
In addition to having similar life histories, the three anguillid spe-
cies underwent a dramatic decline which star ted in the late 1970s
(Dekker & Casselman, 2014; Dekker et al., 2003). The collapse be-
came evident through the analyses of the recruitment indices of
the three species (Figure 3). Recruitment series of glass eels are the
most reliable indices to estimate trends in eel populations through-
out their range, because they are less influenced by local conditions
than indices using older stages. In Europe, recruitment has de-
creased by 90%–99% since the 1980s (ICES, 2015). Concerning the
American eel, the recruitment in the upper Saint Lawrence River and
Lake Ontario has nearly ceased in one of the most productive areas
(Casselman, 2003) and commercial silver eel CPUE dropped by at
least 50% in 40 years in the Saint Lawrence River (de Lafontaine,
Gagnon, & Côté, 2009). Recruitment of Japanese eel has followed
a very similar decline, corresponding to 80% in the last dec ades
(Dekker et al., 2003; Tanaka, 2014). As a result, IUCN classified
Europea n eel as critically e ndangered in 20 08 (confirmed in 2010 and
2014) (Jacoby & Gollock, 2014a) while American and Japanese eels
have been classified as endangered since 2014 (Jacoby, Casselman,
DeLucia, Hammerson, & Gollock, 2014; Jacoby & Gollock, 2014b).
Meanwhile, conser vation regulations have flourished for the three
species (Figure 4). The European Commission implemented a regu-
lation in 20 07 (Council Regulation [EC] No 1100/2007) establishing
measures for the recovery of the stock of European eel and calling
for a reduction in anthropogenic mortalities. The North American
eel is “endangered” since 2008 as under Ontario’s Endangered
Species Act, which prohibits their fishing and trading. The Canadian
FIGURE3 Recruitment series for the three temperate eel
species. European eel series (black solid line) corresponds to the
Elsewhere Europe index provided by ICES (2015). American eel
recruitment (grey solid line) corresponds to the recruitment in Lake
Ontario through monitoring of eel passage at Moses Saunders
hydroelec tric dam (A. Mathers, Ontario Ministr y of Natural
Resources, personal communication). Japanese eel recruitment
(black dot ted line) corresponds to Japanese catch statistics (data
may include young yellow eels larger than glass eels during 1957–
1977—provided by Statistics Department, Ministry of Agriculture,
Forestr y and Fisheries, Japan, till 2002 and from Fisheries A gency,
Japan since 2003). Data were smoothed using a 5- year moving
geometric mean and expressed as a percentage of the 1960s–1970s
1960 1980 2000
DROUINE AU Et Al.
federal Government is currently considering whether the American
eel should be listed as “Threatened” under the federal Species at
Risk Act. The Japanese eel is “endangered” on the Japanese Red List
published by the Ministry of Environment, Japan in 2013.
The eel population declines possibly result from oceanic changes
(Castonguay, Hodson, Moriarty, Drinkwater, & Jessop, 1994), over-
fishing (Dekker, 2003c; Haro et al., 2000; Tsukamoto, Aoyama, &
Miller, 2003), contamination (Belpaire, Geeraerts, Evans, Ciccotti,
& Poole, 2011; Belpaire, Pujolar, Geeraerts, & Maes, 2016), parasit-
ism (Feunteun, 2002; Kirk, 20 03) and blockage due to dams (Kettle,
Asbjørn Vøllestad, & Wibig, 2011; Moriarty & Dekker, 1997). These
suspected causes match the above- mentioned components of global
change: oceanic modifications resulting from global warming, over-
fishing corresponds to overharvesting of natural resources, the eel
swimbladder parasite is an alien species, blockage due to dams is an
example of habitat loss due to land use, and contamination is the
direct result of the increased pollutant load. The respective ef fect s
of anthropogenic pressures are difficult to disentangle and proba-
bly acted synergistically in the declines of eel populations (Jacoby
et al., 2015; Miller et al., 2016). Here, we focus on the timelines of
events and on the spatial dimension to show how some anthropo-
genic pressures have had a greater impact on certain habitat s (and
consequently eels in these habitats) than others. At last, we discuss
why the overall decline may be interpreted as the result of the com-
bined effect of the global change components and how the cumula-
tive pressures have had a more drastic impact, than 50 million years
of evolution, by radically reducing adaptive capacities of eels. Even
though there are growing concerns about the situation of tropic al
eels and many similarities with temperate eels (Jacoby et al., 2015),
we restricted our analysis to temperate eels of the genus Anguilla be-
cause they are in the worst situation according to IUCN criteria, and
are the three most commercially impor tant species (Jacoby et al.,
2 | COMPONENT 1—GLOBAL WARMING
AND OCEAN MODIFICATION: IMPACTS ON
2.1 | Effect on leptocephalus drift and survival
The decline in eel populations is possibly linked to modifications of
physical conditions in the oceans (Bonhommeau, Chassot, Planque,
et al., 2008; Castonguay, Hodson, Moriarty, et al., 1994; Knights,
White, & Naismith, 1996; Miller et al., 2009, 2016). The synchronous
declines of the three eel species may indicate the involvement of
large- scale drivers, such as changes in oceanic conditions that af fect
hatching and subsequent survival of lar vae (Bonhommeau, Chassot,
Planque, et al., 2008; Castonguay, Hodson, Moriar ty, et al., 1994).
Global warming has had a measurable impact on sea surface temper-
atures (Figure 5) and on different oceanic features (North Atlantic
Oscillation, El Niño Southern Oscillation and North Equatorial
Current) influencing recruitment success of temperate eels. Global
FIGURE4 Timelines of main events
with respect to the five global change
components and management of eel
populations. Committee on the Status
of Endangered Wildlife in Canada
(COSEWIC) is a committee of experts
that assesses and designates which
wildlife species are in some danger of
disappearing from Canada [Colour figure
can be viewed at wileyonlinelibrary.com]
DROUINEAU Et Al .
warming has also had visible impacts on planktonic communities:
important shif ts in the diversity, abundance and spatial distribution
of planktonic species in the Atlantic Ocean, where plankton is cru-
cial for eel larval growth and survival (Beaugrand, 2004; Beaugrand,
Luczak, & Edwards, 2009; Goberville, Beaugrand, & Edwards, 2014).
Indeed, there are correlations between glass eel recruitment
and different oceanic indicators, such as the North Oscillation
Index (NAO) and sea surface temperatures (Table 1). Because
most studies computed statistical correlations, the underly-
ing mechanisms are speculative. Three main mechanisms have
been proposed: a limitation in trophic conditions (Bonhommeau,
Chassot, & Rivot, 20 08; Bonhommeau et al., 2009; Desaunay &
Guerault, 1997; Friedland, Miller, & Knights, 2007; ICES, 2001;
Kettle & Haines, 20 06; Knights, 20 03; Munk et al., 2010), changes
in oceanic currents modifying larval transport (Castonguay,
Hodson, Moriarty, et al., 1994; Friedland et al., 2007; ICES, 2001;
Knights, 2003; Zenimoto et al., 2009), and/or spatial oscillations of
a salinit y front used by adult eels to detect the spawning grounds
which then lead to oscillations in the success of larval transport
(Kimura, Inoue, & Sugimoto, 20 01; Kimura & Tsukamoto, 2006).
In addition to statistical correlations, Lagrangian simulations of
larval drift have also been c arried out to explore some mecha-
nisms (Bonhommeau, Castonguay, Rivot, Sabatié, & Le Pape, 2010;
Bonhommeau et al., 2009; Kettle & Haines, 2006; Kim et al., 2007;
Melià et al., 2013; Pacariz, Westerberg, & Björk, 2014; Zenimoto
et al., 2009). These simulations suggested that, while for European
eel the correlation between NAO and recruitment more likely re-
flect s an indirect effect of trophic conditions in the Sargasso Sea
(Bonhommeau et al., 2009; Pacariz et al., 2014), changes in oce-
anic currents directly af fect Japanese eel larval drift (Kim et al.,
2007; Zenimoto et al., 2009).
Analysing simultaneously the declines of the three species,
Bonhommeau, Chassot, Planque, et al. (2008) highlighted the
synchrony between regime shifts in the Atlantic and Pacific sea
surface temperature, primary production and recruitment of the
three species. They postulated that an increase in sea surface tem-
perature due to climate change led to a higher stratification of the
Sargasso Sea and consequently to a lower primary production,
which could translate into lower food availability for the lepto-
cephali. In recent times, Miller et al. (2016) proposed a more pre-
cise mechanism: The regime shif t resulted in a lower abundance of
diatoms and a higher abundance of cyanobacteria, which may have
resulted in a lower production of carbohydrates which are crucial
for the production of “marine snow,” the main food of eel larvae
(Riemann et al., 2010).
2.2 | Effect on silver eel spawning migration
Climate change can also affect the later stages of eel. Oceanic con-
ditions and climate change can, indirectly, influence river discharge
(Arnell, 1999; Milly, Dunne, & Vecchia, 2005) through modifications
of precipitation regimes (Kettle et al., 2011). The discharge regime
is also modified by water extraction for human use, agriculture
and other industrial processes (Postel & Richter, 2003; Verreault,
Mingelbier, & Dumont, 2012). River discharge and rainfall are im-
portant triggers (direc t or indirect) of silver eel migration (Acou,
Laffaille, Legault, & Feunteun, 20 08; Boubée et al., 2001; Bruijs &
Durif, 2009; Drouineau et al., 2017; Durif & Elie, 2008; Reckordt,
Ubl, Wagner, Frankowski, & Dorow, 2014; Trancart, Acou, Oliveira,
& Feunteun, 2013). Higher river discharge increases migration speed
(Tesch, 2003; Vøllestad et al., 1986). Reduced discharge delays mi-
gration, and eels may even be stopped until the following year if en-
vironmental conditions are not favourable (Drouineau et al., 2017;
Durif, Elie, Gosset, Rives, & Travade, 2003). At last, reduced dis-
charge can lead to higher proportions of eels going through turbines,
since at low flow a higher proportion of water is guided through
the turbines, leading to higher mortalities (Bau et al., 2013; Jansen,
Winter, Bruijs, & Polman, 2007).
3 | COMPONENT 2—INCREASED
CONTAMINATION LOAD: CONTAMINATION
OF EELS AND CONSEQUENCES ON ITS
Eels are vulnerable to contamination because of their high trophic
level and high level of lipid storage (Belpaire et al., 2011; Geeraerts &
Belpaire, 2009; ICES, 2016). Cont aminants that have been found in
eels include the following: organic contaminants (Bilau et al., 2007;
Blanchet- Letrouvé et al., 2014; Guhl, Stürenberg, & Santora, 2014;
Hodson et al., 1994; Kammann et al., 2014; Ohji, Harino, & Arai,
2006), heavy metals (Maes et al., 2005; Nunes et al., 2014; Pannetier
et al., 2016; Pierron, Baudrimont, Dufour, et al., 2008; Pierron,
Baudrimont, Lucia et al., 2008; Yang & Chen, 1996) and pesticides
(Byer et al., 2013; Couillard, Hodson, & Castonguay, 1997; Gimeno,
Ferrando, Sanchez, Gimeno, & Andreu, 1995; Hodson et al., 1994;
Privitera, Aarestrup, & Moore, 2014).
FIGURE5 Ocean temperature anomalies (left panel). Source
(Morice, Kennedy, Rayner, & Jones, 2012; Steffen et al., 2015)
1920 1950 1980 2010
emperature anomaly (°C)
DROUINE AU Et Al.
Therefore, eels are sometimes used as bioindicators of contam-
ination (Amiard- Triquet, Amiard, Andersen, Elie, & Metayer, 1987;
Belpaire & Goemans, 2007; Linde, Arribas, Sanchez- Galan, & Garcia-
Vazquez, 1996; McHugh et al., 2010). Contamination levels are often
above human consumption standards (Bilau et al., 2007; Byer et al.,
2013; Geeraerts & Belpaire, 2009; ICES, 2014, 2016) and have led to
fishing prohibitions in various sites in European countries (Germany,
Belgium, Netherlands, France, Italy) (Belpaire et al., 2016).
These contaminants are widely found in freshwater fishes (Streit,
1998), and their ef fects on fish biology (Fonseca et al., 2014; Gilliers
et al., 2006; Kerambrun et al., 2012) and the danger for human
consumption have been demonstrated (Halldorsson, Meltzer,
Thorsdottir, Knudsen, & Olsen, 2007; Järup, 2003; Schuhmacher,
Batiste, Bosque, Domingo, & Corbella, 1994). While metallic con-
taminants have a long history in countries with ex traction activi-
ties, organic contamination, pesticides and nutrients loads are much
more recent (Malmqvist & Rundle, 2002; Morée, Beusen, Bouwman,
& Willems, 2013). Many of them appeared in the second half of the
twentieth century in relation to agriculture intensification, urbaniza-
tion and industrial activities. During this period, fertilizer utilization
grew exponentially (Figure 6). PCBs and DDT production peaked
around the 1960s (Harrad et al., 1994; Van Metre, Wilson, Callender,
& Fuller, 1998); that is, about 20 years before eels star ted to decline,
and concentrations remain high in river sediments explaining why
levels are still high in eels (ICES, 2016). Persistent organic pollutants
(POPs) had a dramatic ef fect on lake trout (Salvelinus namaycush)
TABLE1 Main references exploring the impact of oceanic conditions on recruitment
Reference Species Oceanic index Proposed mechanisms
Desaunay and Guerault
European eel Oceanic temperature Food availability
Kimura et al. (2001) Japanese eel Southern Oscillation Index
El Niño/Southern Oscillation
Oscillations of the salinit y front that affects
larvae growth and survival during their
ICES (2001) American and European eels North Atlantic
Changes of t ransport due to modification of
Gulf Stream path
Trophic limitations due to oscillation in
Knights (2003) American and European eels North Atlantic
Sea surface temperature
Changes of t ransport due to modification of
Gulf Stream path
Trophic limitations due to oscillation in
Kettle and Haines (2006) European eel Lagrangian circulation model Food availability
Kimura and Tsukamoto
Japanese eel Field observation on salinity
Oscillations of spawning loc ation due to
movement s of salinity front induced by El
Friedland et al. (20 07) European eel and presumably
North Atlantic Oscillation Food availability in the Sargasso Sea larval drif t
Kim et al. (2007) Japanese eel Lagrangian circulation model Success of lar val transpor t due to oscillation of
the North Equatorial Current
and Rivot (2008)
European eel Sea surface temperature in the
Bonhommeau et al.
European eel Lagrangian circulation model
North Atlantic Oscillation Index
Gulf Stream Index
Oscillations biological production in the
Zenimoto et al. (200 9) Japanese eel Lagrangian circulation model Success of larval transport due to oscillation of
the North Equatorial Current
Munk et al. (2010) European eel and presumably
Field observations of oceanic
fronts in the Sargasso Sea
Oscillations of fronts that alter the efficiency
of retention on feeding grounds
Durif, Gjøsæter, and
Vøll es ta d (2 011)
European eel Analysis of a 100- year old time
series of eel abundance
Relationship to NAO and temper ature
conditions in the Sargasso Sea
Pacariz et al. (2014) European eel Lagrangian circulation model Decline of success of larval t ransport due to
current modifications (rejected)
Miller et al. (2016) European, Japanese and
Field measurement of diatoms
and cyanobacterial abundances
in the Sargasso Sea
Lower availability of food after oceanic regime
DROUINEAU Et Al .
in Lake Ontario (Cook et al., 2003) and concentrations peaked in
American eel in Lake Ontario in the late 1960s, about 20 years be-
fore the American eel recruitment collapse (Byer et al., 2015). The
increased nutrient load to water bodies has caused detrimental im-
pacts on humans and aquatic ecosystem health (Grizzetti, Bouraoui,
& Aloe, 2012; Grizzetti et al., 2011) and continued to increase until
the mid- 1990s before declining in many rivers (Minaudo, Meybeck,
Moatar, Gassama, & Curie, 2015). Fertilizers are still used at a very
high level (Figure 6). Moreover, new contaminants are appearing in
the water and in eels, such as perfluorooctanesulfonic acid, textile
dyes, musk compounds, perfluorinated substances, organophospho-
rus flame retardants and plasticizers (ICES, 2016).
There have been a few cases of direct eel mor talities due to
contaminants (Dutil, 1984; Dutil, Besner, & McCormick, 1987), but
in the majority of cases, the impact is at the sublethal level ranging
from tissue damage, stress, effects on osmoregulation, behaviour
alteration, hormonal per turbation and genotoxic effects (Couillard
et al., 1997; Geeraerts & Belpaire, 2009). Cont aminants may also be
transferred to the offspring resulting in larval malformation (Byer
et al., 2013; Foekema, Kotterman, de Vries, & Murk, 2016; Rigaud
et al., 2016; Robinet & Feunteun, 2002). A s a fatt y fish, eels are par-
ticularly sensitive to contamination. Most contaminants are highly
concentrated in the lipid stores (Robinet & Feunteun, 20 02) and af-
fect lipid metabolism (Corsi et al., 2005; Fernández- Vega, Sancho,
Ferrando, & Andreu- Moliner, 1999; Pierron et al., 2007). This is es-
pecially critical at the silver eel stage when lipid levels are highest
(over 13%) to achieve their transoceanic migration to the spawn-
ing grounds (Belpaire et al., 2009; Van Den Thillart, Palstra, & Van
Ginneken, 2007; Van Den Thillart et al., 20 04; van Ginneken & van
den Thillart, 2000). For female eels, 67% of their fat store is spent
on the spawning migration and oocyte maturation (Palstra & van
den Thillart, 2010). As lipids are mobilized during spawning migra-
tion, contaminant s are more likely to be released into the blood at
high concentrations, thus negatively af fecting gonad maturation and
oocyte produc tion, as they do in other fish species (Baillon et al.,
2015; ICES, 2016; Pierron et al., 2014), and also impairing migration
success (Geeraer ts & Belpaire, 20 09; Pierron, Baudrimont, Dufour,
et al., 2008; Robinet & Feunteun, 20 02). As a summary, contami-
nants can act as a classical stressor during the continental stage of
eel, but then have the potential to dramatically impair maturation
and migration success, that is the whole reproduction success.
4 | COMPONENT 3—FRAGMENTATION
AND HABITAT LOSS: FRAGMENTATION BY
WEIRS AND DAMS AND CONSEQUENCES
ON UPSTREAM AND DOWNSTREAM
4.1 | Movements, habitats and fragmentation
Movement is a key feature of living organisms to find food, mates
and avoid predation (Nathan et al., 2008). Several types of move-
ments can be distinguished. The first type, called “station keeping”
(Dingle, 1996), takes place within the home range of the animal and
corresponds to simple movements for foraging and predation avoid-
ance. The t wo other t ypes of movement, ranging and migration,
occur outside the home range (Dingle & Drake, 2007). Ranging is
dedicated to the search for a specific resource (mate, food, etc.) and
stops when the resource is found (Jeltsch et al., 2013). Migration is
generally triggered by physiological and environmental cues and not
by the search for a specific resource such as food or mates. It affec ts
most individuals in the population, occurs over a long timescale,
requires orientation and suggests a return journey (Dingle, 1996;
Dingle & Drake, 2007).
Diadromous fish, such as eels, undergo two long migrations
(Tesch, 2003): The first migration, from the spawning grounds to
their grow th habitat, includes a phase of active upstream migra-
tion in river catchments during the early years of their continen-
tal life- stage (Castonguay, Hodson, Couillard, et al., 1994; Fukuda,
Aoyama, Yokouchi, & Tsukamoto, 2016; Imbert, Labonne, Rigaud,
& Lambert, 2010). During the second migration, eels return to the
oceanic spawning grounds from their growth habitats in rivers or
coastal waters. Eels may also move between different habitats
during their continental stage (Arai & Chino, 2012; Béguer- Pon,
Castonguay, Benchetrit, et al., 2015; Daverat, Tomas, Lahaye,
Palmer, & Elie, 20 05; Kaifu, Tamura, Aoyama, & Tsukamoto, 2010;
Yokouchi et al., 2012), movements which correspond to station
keeping and ranging.
The construction of dams accelerated worldwide during the
1950/1960s (Dynesius & Nilsson, 1994; MacGregor et al., 2009;
Postel & Richter, 2003) (Figure 7), about 20 years before the eel pop-
ulation declined. This massive construction of dams has restrained
eel movements and available habitats. The construction of hydro-
power dams during the t wentieth century in the St. Lawrence catch-
ment caused a 40% habitat loss for the North American eel in this
basin (Verreault, Dumont, & Mailhot, 2004). The situation is similar
FIGURE6 Global fertilizer consumption in OECD countries
(grey) and in the world (black). Source (Steffen et al., 2015,
International Fertilizer Industry Association Database)
1920 1950 1980 2010
ilizer consumption (million tonnes
DROUINE AU Et Al.
or worse in the United States (Busch, Lar y, Castilione, & McDonald,
1998), especially because most dams lack fishways (MacGregor
et al., 2009). In Europe, 50%–90% of habitats were lost by the end
of the twentieth century (Feunteun, 2002). For the Japanese eel, ap-
proximately 75% of effective habitats were lost between 1970 and
2010 in Japan, Korea, Taiwan and China, with a maximum in China
(>80%) and Taiwan (~50%) (Chen, Huang, & Han, 2014).
Intensive dam constructions in Spain, Morocco and Portugal,
have had dras tic consequence s on European eel distribution (Clavero
& Hermoso, 2015; Lobon- Cervia, 1999; Nicola, Elvira, & Almodóvar,
1996), possibly affecting the sex ratio as this area yields mainly male
eels and is closest to the spawning area (Kettle et al., 2011).
Dams and weirs are not the only fac tors affecting eel habitats:
Rivers provide multiple goods and services to society (Elliott &
Whitfield, 2011; Postel & Richter, 2003; Wolanski, McLusky, van den
Belt, & Costanza, 2011) that have led to river channelization, hydro-
morphological modifications, drying out of lateral wetland, wetland
drainage, water extraction, modification of land use in the floodplain
that can lead to higher erosion and sedimentation (Basset et al.,
2013; Elliott & Hemingway, 2002; Postel & Richter, 2003). As an ex-
ample, typical eel habit ats, such as estuarine marshes and intertidal
zones, have been lost because of flood protection walls, agriculture
activities and navigation (Gros & Prouzet, 2014). In Japan, catch re-
duction rates in several rivers and lakes were positively correlated
with the rate of revetment along rivers and around lakes (Itakura,
Kitagawa, Miller, & Kimura, 2015), and also, the condition factor of
eels and prey diversity were significantly lower in these modified
habitat s (Itakura, Kaino, Miyake, Kitagawa, & Kimura, 2015).
4.2 | Blockage during upstream migrations
During their first year in continental waters, eels display an ac-
tive migratory behaviour and then shif t to a resident behaviour
(Benchetrit et al., 2017; Imber t et al., 2010). Resident behaviour does
not exclude habitat shifts (Daverat & Tomás, 2006) although these
types of movement correspond more to ranging than strict migra-
tion (Dingle & Drake, 2007). Upstream migration has a cost and it s
evolutionary benefit is still unclear as eels can settle in a wide range
of habitat s (Daverat et al., 2006; Marohn, Jakob, & Hanel, 2013;
Tsukamoto, Nakai, & Tesch, 1998; Yokouchi et al., 2012). Glass eels
with high feeding rate and fast weight gain have a higher propen-
sity to migrate (Bureau du Colombier, Lambert, & Bardonnet, 2008).
These glass eels also display a more gregarious and less aggressive
behaviour (Geffroy & Bardonnet, 2012). Habitat selection could be
a trade- off between growth (generally higher in downstream habi-
tats), sur vival (generally higher in upstream habitats), competition
avoidance (higher competition in downstream habitats) and ener-
getic cost of migration (Drouineau et al., 2014; Edeline, 2007; Mateo,
Lambert, Tétard, Castonguay, et al., 2017). Habitat selection is also
partly related to genetic or epigenetic polymorphism (Côté et al.,
2014; Gagnaire et al., 2012; Mateo, Lambert, Tétard, Castonguay,
et al., 2017; Pavey et al., 2015; Podgorniak, Milan, et al., 2015). In
such a scheme, habitat selection would be the result of a fitness
optimization process in which fitness in a habitat would depend on
habitat characteristics, competition in the habit at, but also individual
variability of grow th rates due to the existence of genetically distinct
clusters of individuals (Côté et al., 2015; Mateo, Lambert, Tétard,
Castonguay, et al., 2017).
Given this plasticity in habitat use, the consequences of obsta-
cles on upstream migrations are difficult to assess. Methods have
been proposed to assess the passability of obstacles (Briand, Fatin,
Feunteun, & Fontenelle, 2005; Drouineau et al., 2015; Tremblay,
Cossette, Dutil, Verreault, & Dumont, 2016). Densities of eels are
higher downstream of obstacles. This leads to (a) increased com-
petition between individuals, which can subsequently result in
lower survival (Bevacqua, Melià, de Leo, & Gatto, 2011; Vøllestad &
Jonsson, 1988), (b) increased susceptibility to predation (Agostinho,
Agostinho, Pelicice, & Marques, 2012; Drouineau et al., 2015; Garcia
De Leaniz, 2008; Larinier, 2001) and overfishing (Briand et al., 2005;
Dekker, 2003c), and (c) possible modification to the sex ratio, as sex
determination is density- dependent (Davey & Jellyman, 2005; De
Leo & Gatto, 1996; Poole, Reynolds, & Moriarty, 1990; Roncarati,
Melotti, Mordenti, & Gennari, 1997; Tesch, 2003).
At last, obstacles to upstream migration can act as a permanent
selection pressure (Mateo, L ambert, Tétard, & Drouineau, 2017;
Podgorniak, Angelini, et al., 2015; Podgorniak, Milan, et al., 2015).
Côté et al. (2014) demonstrated the existence of two clusters of in-
dividual eels with differing genetic basis: a cluster of slow growers
and a cluster of fast growers, while Pavey et al. (2015) demonstrated
the existence of genetically distinct ecotypes, with different growth
rates and different sex ratios. By impairing migration within catch-
ments, obstacles can decrease the fitness of some types of individ-
uals; those individuals who genetically belong to the “freshwater
habitat ” may not be able to reach suitable habitat s or will suffer dam-
age during their downstream migration (Mateo, Lambert , Tétard, &
FIGURE7 Accumulative number of large dams in OECD
countries (grey) and in the world (black). Source (Steffen et al.,
2015; World Commission on Dams, 2000)
1920 1950 1980 2010
DROUINEAU Et Al .
4.3 | Impaired downstream migrations
Most studies dealing with downstream migration have focused on
mortality due to passage through hydropower turbines (Boubée &
Williams, 2006; C alles et al., 2010; Carr & Whoriskey, 2008; Coutant
& Whitney, 2000; Gosset, Travade, Durif, Rives, & Elie, 20 05;
Pedersen et al., 2012; Winter, Jansen, & Bruijs, 2006). Several fac-
tors influence the mortality induced by hydropower plants:
1. Turbine characteristics: The mortality due to strikes by Kaplan
turbines is generally greater than 15% and sometimes as high
as 100% depending on fish length, wheel diameter, nominal
discharge flow and speed of rotation (Gomes & Larinier, 2008).
For Francis turbines, Calles et al. (2010) estimated a mortality
rate of 60% at a Swedish site while a mortality rate of about
16% was found at an American site (Richkus & Dixon, 2003).
Even if they sur vive passage through the turbines, eels can
be wounded and have a reduced chance of reaching the spawn-
2. Site configuration can greatly influence the probability that a fish
will pass through or by-pass the turbines. As silver eels follow the
main flow (Jansen et al., 2007), the orientation of the water intake
with respect to the main channel influences the probability of tur-
bine passage (Bau et al., 2013). Different types of barriers have
been proposed to divert eels from turbine passage, such as fish-
friendly trashracks (Raynal, Chatellier, Courret, Larinier, & David,
2014; Raynal, Courret, Chatellier, Larinier, & David, 2013), flow
field manipulation (Piper et al., 2015), light (Hadderingh, Van Der
Stoep, & Hagraken, 1992; Patrick, Sheehan, & Sim, 1982) and in-
frasound barriers (Sand, Enger, Karlsen, Knudsen, & Kvernstuen,
2000; Sand et al., 2001). The installation of bypasses is also a miti-
gation measure to prevent passages through turbines (Durif et al.,
2003; Gosset et al., 2005; Haro, Watten, & Noreika, 2016).
3. Environmental conditions: In a period of low discharge, when the
flow through the turbine is high compared to the flow over weir,
more eels will pass through the turbines than at high discharge,
when the turbine flow is small compared to the weir flow.
4. Obstacle location within the catchment: As eels are not uniformly
distributed within a river catchment (Ibbotson, Smith, Scarlett, &
Aprhamian, 2002), the number of eels impacted by a given facility
depends on the number of eels that settle upstream the facility.
Therefore, it is necessary to estimate the distribution of fish
within catchments to assess the effect of hydropower plants at
the catchment scale. In the SE AHOPE model, the total mortality
induced by hydropower plants in a given catchment was esti-
mated by coupling a model that predicts the proportion of fish
killed when passing each individual plant with a model that pre-
dicts the spatial distribution of eels within the catchment (Jouanin
et al., 2012).
However, direct m ortality is not the only im pact obstacles can have
on downstream migrants. First, sublethal injuries can occur during ob-
stacle passages because of impingements on hard structures (even in
the absence of turbines), which can then impair spawning migration
success (Bruijs & Durif, 2009). Predation during downstream passage
has also been recorded for many fish species (Garcia De Leaniz, 20 08;
Muir, Marsh, Sandford, Smith, & Williams, 2006; Williams, Smith, &
Muir, 2001). Moreover, increased energy costs induced by obstacle
passage may have a delayed impact on migration success and fecun-
dity: Silver eels stop feeding during the spawning migration, and their
lipid stores are crucial to achieve the oceanic migration and produce
oocytes (Van Ginneken & van den Thillart, 20 00). Delays induced by
obstacles can impair escapement, especially when the environmen-
tal migration suitability window is limited (Drouineau et al., 2017;
Verbiest, Breukelaar, Ovidio, Philippart, & Belpaire, 2012). At last,
similarly to obstacles to upstream migration, obstacles to downstream
migration affect specific types of individual: individuals that settle up-
stream of the obstacle (i.e. individuals that settle in upstream habitats
and individuals that were able to pass the obstacle), as such, obstacles
may have the potential to exer t a selection pressure on the population
(Mateo, Lambert, Tétard, & Drouineau, 2017).
5 | COMPONENT 4—ALIEN SPECIES:
EFFECTS OF ALIEN PARASITOID
Although competition is possible with some alien species such as
the European catfish (Bevacqua et al., 2011), or even with intro-
duced American (Han et al., 20 02) and European eels in East Asia
(Aoyama et al., 2000), A . crassus is the alien species that has the most
documented and widespread impact on eels, at least for European
and American eels. A. crassus is a natural parasite of Japanese eel
which was introduced into Europe in the mid- 1970s, early 1980s,
probably through the aquaculture trade (Koops & Hartmann, 1989).
It is now widespread in Europe (Becerra- Jurado et al., 2014; Evans
& Matthews, 1999; Kennedy & Fitch, 1990; Kirk, 20 03; Lefebvre,
Contournet, Priour, Soulas, & Crivelli, 20 02; Neto, Costa, Costa, &
Domingos, 2010; Norton, Rollinson, & Lewis, 2005) and Northern
Africa (Dhaouadi et al., 2014; El Hilali, Yahyaoui, Sadak, Maachi,
& Taghy, 1996; Hizem Habbechi, Kraiem, & Elie, 2012; Koops &
Hartmann, 1989; Maamouri, Gargouri, Ould Daddah, & Bouix, 1999).
Systematic monitoring of eel diseases is still limited to a few coun-
tries, impairing our abilit y to assess the overall prevalence (ICES,
2015). However, many studies have reported a significant preva-
lence at sites in both Nor th America and Europe (Aieta & Oliveira,
2009; Becerra- Jurado et al., 2014; Denny, Denny, & Paul, 2013) and
an analysis of the European Eel Quality Database confirmed the
prevalence of the infection in Europe (Belpaire et al., 2011).
The invasion in North America started for the same reason, a
few years af ter its introduc tion into Europe. The first record oc-
curred in the second half of the 1990s in Texas (Fries, Williams, &
Johnson, 1996) and then in Chesapeake Bay and the Hudson River
(Barse & Secor, 1999). The invasion then quickly spread in the United
States and in Canada (Aieta & Oliveira, 2009; Denny et al., 2013;
Hein, Arnott, Roumillat, Allen, & de Buron, 2014; Machut & Limburg,
DROUINE AU Et Al.
2008; Rockwell, Jones, & Cone, 2009). Although transmission is
possible in brackish waters (Kirk, Kennedy, & Lewis, 20 00; Kirk,
Lewis, & Kennedy, 2000; Lefebvre et al., 2002; Reimer, Hildebrand,
Scharberth, & Walter, 1994), the level of infection is lower than in
freshwater (Kirk, 2003; Kirk, Kennedy, et al., 2000; Kirk, Lewis, et al.,
This swimbladder parasite has multiple impacts on its host. The
parasite causes inflammation of the swimbladder leading to mul-
tiple bac terial infections, stress and loss of appetite (Kirk, 2003;
Lefebvre, Fazio, Mounaix, & Crivelli, 2013). However, the most se-
rious damage is on the swimbladder itself. The infection may alter
the gas composition of the swimbladder, block the pneumatic duct,
impairing the organ’s function (Kirk, Lewis, et al., 2000; Lefebvre
et al., 2013) leading to necrosis in the most ex treme cases (Molnár,
Székely, & Perényi, 1994; Würtz & Taraschewski, 2000). The alter-
ation of the swimbladder has a direct impact on swimming capac-
ity (Sprengel & Lüchtenberg, 1991; Székely, Palstra, Molnár, & van
den Thillart, 2009). It may imperil the transoceanic spawning migra-
tion (Clevestam, Ogonowski, Sjoberg, & Wickstrom, 2011; Palstra,
Heppener, Van Ginneken, Székely, & Van den Thillart, 2007), espe-
cially because migrant eels display important diurnal vertical migra-
tions (Béguer- Pon, Castonguay, Shan, et al., 2015; Chow et al., 2015;
Righton et al., 2016) that require buoyancy control. This higher
energetic cost of migration, due to a malfunctioning swimbladder,
will affect individuals which may already have reduced lipid storage
available, due to the infection (Marohn et al., 2013).
6 | COMPONENT 5—EXPLOITATION OF
NATURAL RESOURCES: AN INTENSIVE
EXPLOITATION OF EELS AT ALL THEIR
STAG E S
Eels are targeted by recreational and commercial fisheries at all con-
tinental life stages (glass eels, yellow eels and silver eels) with a great
variety of active and passive gears (Haro et al., 2000; Tesch, 2003).
Yellow and silver eels have been exploited for a long time as attested
by representations of eels in prehistoric pictographs (Citerne, 1998,
2004). Eel was an important food resource for Native Americans
(MacGregor et al., 2009) and is a traditional food in Japan and East
Asia (Tatsukawa, 2003). The first official record of European eel
fisheries dates back to 1086 (Dekker & Beaulaton, 2016). In con-
trast to the situation for many commercial species, the culture of
eel is not a closed system in that it is still dependent on wild- caught
glass eels. Artificial reproduction and rearing of glass eels have
only been achieved for the Japanese eel (Kagawa, Tanaka, Ohta,
Unuma, & Nomura, 20 05; Tanaka, Kagawa, & Ohta, 2001; Tanaka,
Kagawa, Ohta, Unuma, & Nomura, 2003) although these operations
are still not commercially viable (Okamura, Horie, Mikawa, Yamada,
& Tsukamoto, 2014). Artificial reproduction has been achieved
in European (Palstra & van den Thillart, 2009) and American eels
(Oliveira & Hable, 2010) but not rearing of glass eels.
The main shift in the traditional artisanal eel fisheries occurred
as a result of the demand from on- growing aquaculture (Haro et al.,
2000; Moriarty & Dekker, 1997). According to FAO statistic s, eel
farming is now responsible for 90% of total eel production (vs wild-
caught eels) and Japan is thought to consume 70% of tot al freshwa-
ter eel production (Shiraishi & Crook, 2015). While eel aquaculture
started in the late nineteenth centur y and early twentieth century
in eastern Asia, it turned into a stable industry after World War II
(Ringuet, Muto, & Raymakers, 2002). The high value of eel in Eastern
Asia food markets led to the development of highly competitive
aquaculture farms (Lee, Chen, Lee, & Liao, 2003; Liao, 2001). The
development of intensive farming explains why despite the decline
in the wild population, the production of Anguilla spp. increased
nearly 20- fold between 1950 and 2007 (Crook & Nakamura, 2013).
As these farms depend on wild- caught animals, the demand for glass
eel increased considerably and prices climbed to very high levels,
completely transforming the industry. The shortage of Japanese
glass eels since the early 1970s leads aquaculture farms to import
European and American glass eels (Haro et al., 2000; Lee et al.,
2003; Moriarty & Dekker, 1997; Ringuet et al., 2002), leading to an
increase in fishing effort in Europe and a peak in landings in 1976
(Briand, Bonhommeau, Castelnaud, & Beaulaton, 2008) and to
the development of a large fisher y targeting glass eels from North
America (Meister & Flagg, 1997). After a period of less favourable
market conditions, the prices soared again during the early 1990s
(Briand et al., 2008). A threefold increase in prices of European glass
eel was observed between 1993 and 1997 (Ringuet et al., 2002), re-
sulting in a “gold rush” for entry into the North American fishery
(Haro et al., 200 0). Because of these incredibly high prices, eel be-
came the most valuable species landed in France in the early 2000s
(Castelnaud, 20 00) and Europe exported half of its production to
Asia in the mid- 200 0s (Briand et al., 2008). The increase in fishing ef-
fort led to very high exploit ation rates in certain French and Spanish
catchments (Aranburu, Diaz, & Briand, 2016; Briand et al., 2005;
Prouzet, 2002) as Spain and France recruit the highest proportion of
European eel (Dekker, 2000a). In a similar manner, high exploitation
rates were observed in catchments on Canadian Atlantic seaboard
(Jessop, 2000) or in Taiwan (Tzeng, 1984). In France, about 25% of
the arriving glass eels were harvested by commercial fisheries, and
this estimate did not include the catch from illegal fisheries (Figure 8;
Drouine au, Beaulaton, L ambert, & Briand, 2016). In Japan, these p ro-
portions rose from about 25% in the early 1950s to approximately
40% in the 1980s (Tanaka, 2014). The Eel European Regulation has
limited the fishing effort and required that 60% of caught eels be
dedicated for restocking. Moreover, European eel exports have been
restricted after its inclusion on Appendix II of the Convention on
Trade of Endangered Species in 2009 and a ban of all imports and ex-
port s from and to the European Union implemented in 2010 (Nijman,
2015). In Japan, glass eel fisheries are forbidden and a special licence
is required to capture seed for aquaculture and research. Specific
permission is now required for aquaculture, and restrictions have
also been implemented in China and Taiwan. In 2014, China, Japan,
DROUINEAU Et Al .
the Republic of Korea and Taiwan agreed to restrict “initial input”
into farms of glass eel of Japanese eel.
The catch of silver eels has decreased throughout the world
(Cairns et al., 2014; Dekker, 2003a; Tatsukawa, 2003; Tsukamoto,
Aoyama, & Miller, 2009) because of a reduction in abundance of
the stock and because of a decrease in fishing effort, accelerated
by recent management measures. Silver eel fisheries have, for ex-
ample, completely disappeared in Taiwan (Tzeng, 2016) and are
restricted in 11 prefectures in Japan (Jacoby & Gollock, 2014b). In
Europe, the decline in the silver eel catches has preceded the de-
cline in recruitment (Dekker, 2003a). Silver eel fisheries used to
predominate at the northern edge of their distribution area and in
the western Mediterranean (Aalto et al., 2016; Amilhat, Farrugio,
Lecomte- Finiger, Simon, & Sasal, 20 08; Dekker, 2003b, 20 03c), and
in some catchments, exploitation rates can still be high. Regarding
the American eel, silver eel landings used to be dominated by-
catches in the Saint Lawrence River (Castonguay, Hodson, Couillard,
et al., 1994), but they have also severely declined and a large- scale
licence buyout in Quebec has recently accelerated this trend (Cairns
et al., 2014).
7 | WHEN FORTY YEARS OF GLOBAL
CHANGE HAS HAD A GREATER IMPACT
THAN THE ICE AGES OR CONTINENTAL
The genus Anguilla appeared more than 50 million years ago during
the Eocene ( Tsukamoto & Aoyama, 1998). Japanese eel is thought to
have appeared about 15 million years ago (Lin, Poh, & Tzeng, 2001),
and American and European eels separated about 3 million years ago
during the emergence of the Isle of Panama (Jacobsen et al., 2014).
Those species have sur vived enormous changes: a succession of ice
ages (the las t ice age maximum occur red approximately 2 2,000 years
ago) and continental drift that has progressively increased the dis-
tance between the spawning grounds and growth habitat s (Knight s,
2003). This demonstrates their evolutionar y robustness (Knights,
2003) and remarkable adaptive capacity (Mateo, Lambert, Tétard,
Castonguay, et al., 2017) based on adaptive phenotypic plastic-
ity (Côté et al., 2014; Daverat et al., 2006; Drouineau et al., 2014)
and genetic polymorphism (Gagnaire et al., 2012; Pavey et al., 2015;
Pujolar et al., 2014; Ulrik et al., 2014). Despite millions of years of ad-
aptation, these three eel species have undergone a dramatic decline
in only a few decades.
Identif ying the main drivers of the eel decline is still in debate.
The main arguments to suppor t the importance of specific stressors
are based on the synchrony between the time of the collapse in eel
and the stressor. However, many factors impair our ability to disen-
tangle their respective effects. First, the simultaneous decline of the
three species strongly sug gests the influence of large- scale factors
and therefore of a possible oceanic influence. However, other stress-
ors display very similar increasing trends at the global scale before
the beginning of the decline (4; 6; 7). Moreover, the beginning of the
eel decline is very difficult to identify because of the complex life
cycles of the species (Figure 2) and their long life expectancy (up to
30 years). It would be interesting to compare with tropic al species
that also show signs of decline but comparative data with Southern
hemisphere tropical species are scarce (Jacoby et al., 2015; Jellyman,
2016). Second, robust quantitative historical data on eel and the an-
thropogenic pressures are lacking for this period. Third, where these
data do exist they mainly come from specific river catchment s and
it is not possible to extrapolate these data to the whole distribution
area because the anthropogenic pressure does not have the same
effect every where and eels display a great diversity in life history
traits. Each stressor probably played a role in the collapse and the
combination of stressors in the second half of the twentieth cen-
tury probably had a cumulative effect that heightened the overall
effect of the individual stressors (Jacoby et al., 2015; Miller et al.,
2016). The decline occurred about 30 years after the Second World
War, that is approximately one to three eel generations. This period
corresponds to a period of high economic development “Les Trentes
Glorieuses,” in which agricultural production process, industrial pro-
cess and energy consumption quickly increased. This can be seen
through the acceleration of many indicators since the 1950s/1960s
listed in the study of Steffen et al. (2005), for example world popula-
tion, gross domestic product (increased by a factor of 15 since 1950),
FIGURE8 French glass eel
exploitation rates expressed as the ratio
of catch (tonnes) to recruitment (tonnes).
Catches correspond to an appraisal of
historical catches based upon market and
fisher y data (Briand et al., 2008) while
recruitments were estimated using the
model GEREM (Drouineau, Beaulaton,
et al., 2016)
1980 1990 2000 2010
DROUINE AU Et Al.
world petroleum consumption which has increased by 3.5× since
1960, motor vehicles by a factor of 16 since the early 1950s and in-
creased water use for human consumption and agriculture (Figure 9).
This acceleration of human activity and consumption has been re-
ferred to as the “Great Acceleration” (Steffen, Broadgate, Deutsch,
Gaffney, & Ludwig, 2015; Stef fen et al., 2005), and occurred about
20 years before the first signs of the decline in eel populations, that
is one to two eel generations. As mentioned earlier, river, estuaries
and ecosystems have suffered intense modifications over this period
(Basset et al., 2013; Elliott & Hemingway, 2002; Elliott & Whitfield,
2011; Postel & Richter, 2003, 20 03; Wolanski et al., 2011). Eel pop-
ulations are likely affected by global change as a whole, rather than
by one specific anthropogenic pressure, explaining why Castonguay,
Hodson, Couillard, et al. (1994) could not identif y a primary cause for
the decline of the American eel.
8 | THE RESILIENCE OF EELS SEVERELY
IMPAIRED BY GLOBAL CHANGE
Several factors contribute to the resilience of eel populations. First,
the presence of a brackish/marine contingent (which skip the fresh-
water phase) can buffer the pressures specific to the catadromous
contingen t such as dams, conta mination, fishin g or the parasite (IC ES,
2009). In addition, their ver y large diet spectrum (Sinha & Jones,
1967; Tesch, 2003), their resistance to fluctuations in temperature,
salinit y, oxygen, food availability and temporary emersion (Brusle,
1991; Tesch, 2003) allow them to grow in a very large range of habi-
tats. This plasticity in growth habitat can generate a “storage ef fect ”
and a “portfolio effect” that mitigate against environmental variabil-
ity (ICES, 2009). In a complex life cycle, a storage ef fect corresponds
to a situation where a specific stage of long duration and of limited
sensitivity to environmental conditions, buffers the effects of en-
vironmental conditions on other stages. For eels, the long duration
of the continental growth phase and its variability across habitats
with generational overlaps allows the species to buf fer the faster
cyclic variations of oceanic conditions affecting recruitment (even
in a single cohort, some individuals are likely to face unfavourable
oceanic conditions while others will face more favourable oceanic
conditions during their spawning migration, reproduction and lar-
val drift of their offspring) (Secor, 2015a). A portfolio effect corre-
sponds to the expression “don’t put all your eggs in the same basket.”
For eels, their large adaptive capacity allows them to settle in a wide
range of habitats, smoothing out environmental fluctuations in each
habitat: If one habitat is temporarily unsuitable, it is compensated by
other habitats that remain suitable (Secor, 2015a). More generally,
the large diversity of tactics during the continental phase and pre-
sumably during the spawning migration may correspond to remark-
able bet- hedging well suited to address environmental variability
(Daverat et al., 2006; Righton et al., 2016). The environmental sex
determination may also be a compensatory mechanism: The higher
production of females in a context of depleted population may
FIGURE9 Various indicators of the
Great Acceleration for OECD countries
(grey) or the entire world (black). GDP:
gross domestic product. Carbon dioxide
from firn and ice core records (Law Dome,
Antarc tica) and Cape Grim, Australia
(deseasonalized flask and instrumental
records). Sources (Steffen et al., 2005),
population (Goldewijk, Beusen, & Janssen,
2010), CO2 (Etheridge et al., 1996;
Langenfelds et al., 2011; MacFarling
Meure, 20 04; MacFarling Meure et al.,
2006), water use (Alcamo et al., 2003; aus
der Beek et al., 2010; Flörke et al., 2013),
energy use (GEA Writing Team, 2012), and
GDP (World Bank indicators)
1920 1950 1980 2010
Global population (billion
1920 1950 1980 201
Urban population (billion)
1920 1950 1980 2010
Real GDP (trillion US$)
1920 1950 1980
Carbon dioxide (ppm)
1920 1950 1980 2010
World primary energy use (EJ)
1920 1950 1980 201
Global water use (1000 km3)
DROUINEAU Et Al .
mitigate the reduction in eggs production that would result from the
decline in silver eel abundance (Geffroy & Bardonnet, 2016; Mateo,
Lambert, Tétard, & Drouineau, 2017), especially since eels have a
Then, how might have global change led to such a fast collapse
despite eel adaptive capacities and those compensator y mecha-
nisms? Eels are panmictic and thus have long been considered ge-
netically homogeneous; however, recently a genetic polymorphism
in eel populations was found to be correlated with environmental
gradients (Côté et al., 2014; Gagnaire et al., 2012; Pavey et al., 2015;
Pujolar et al., 2014; Ulrik et al., 2014). These correlations are thought
to result from spatially variable selection (some individuals are ge-
netically more adapted than others to survive in some habitats) or of
genetically based habitat selection (some t ypes of individuals tend to
settle preferentially in some habitats to maximize their fitness). The
existence of genetically distinct types of individuals which are more
or less adapted to the different types of habitats available within
their distribution area (northern vs southern habit ats, marine vs
brackish vs freshwater habitats), that is ecotypes (Pavey et al., 2015),
combined with a large phenotypic plasticity are assumed to play the
main role in eel adaptive capacity, enabling the species to address
the wide environmental heterogeneity at both the distribution and
catchment scale (Drouineau et al., 2014; Mateo, Lambert, Tétard,
Castonguay, et al., 2017). In such a scheme, individuals are able to
grow and sur vive in a wide range of habitats thanks to phenotypic
plasticity but some individuals are more adapted to some habitats
than others (ecotypes), and all individuals reproduce together (pan-
mixia) ensuring that ecotypes are reshuffled in each generation. The
synergy of phenotypic plasticity and genetic polymorphism could
explain how a panmictic population can survive in such a wide and
varied distribution area and be the basis for the adaptive capacities
In this review, we have highlighted that not all pressures affect
all habitats and individuals evenly. Indeed, obst acles af fect mostly
individuals that settle preferentially in upstream habitats and habitat
loss mainly affects males located in the south- western part of the
range of the European eel. Anguillicola crassus has a greater impact
on individuals that settle in freshwater habitats as opposed to estua-
rine or mar ine (Kirk, 2003). At last, f isheries are not uniformly dis trib-
uted, with European silver eel fisheries mainly occurring at the edge
of the distribution area, especially the northern edge, although fish-
eries are also impor tant along the Mediterranean coast (Castonguay,
Hodson, Couillard, et al., 1994; Dekker, 2003b, 2003c; Moriart y &
Dekker, 1997), and glass eel fisheries in the core (Dekker, 2003c). By
affecting different habitats, anthropogenic pressures affect life his-
tory tr aits and ecoty pes in different ways (Figure 10). Climate change
and glass eel fisheries probably affect all ecotypes: Climate change
affects recruitment success. Glass eel fisheries, though not evenly
distributed in the distribution area, generally operate downstream
of river catchments and consequently harvest evenly all incoming
glass eels. On the other hand, all the other anthropogenic pressures
tend to affect ecotypes corresponding to more upstream habitats.
As such, these anthropogenic pressures reduce the fitness of those
individuals and can become an important selective pressure (Mateo,
Lambert, Tétard, Castonguay, et al., 2017; Mateo, Lambert, Tétard, &
Drouineau, 2017). For example, half the American silver eels migrat-
ing down the St. Lawrence River, one of the most productive areas
for American eel (C asselman, 2003), have been killed by hydropower
dams and fisheries (Verreault & Dumont , 2003). Such selection pres-
sure over 30 years or more (one to two eel generations) may have
reduced the prevalence of individuals adapted to such types of hab-
itats (northernmost area, longest migration from spawning grounds)
in the panmictic eel population and explain why recruitment to the
St. Lawrence River has been so much more reduced than elsewhere
in their distribution range. Reducing this genetic polymorphism as a
result of anthropogenic- induced selection may irrevocably alter the
species capacit y to adapt and modif y its sex ratio.
In addition, by decreasing the diversit y of ecotypes and conse-
quently, decreasing the capacity of eels to live in a wide range of
habitat s, anthropogenic pressures may have reduced the portfolio
and storage effects which, as we said before, are crucial to address
environmental variabilit y and to improve resilience. In view of this,
diversit y is crucial for temperate eels (Secor, 2015b) and manage-
ment should preserve this diversity to ensure population resilience.
Moreover, it is crucial to improve our knowledge of the mechanisms
involved in eel adaptation and of the effec ts of anthropogenic pres-
sures on their capacity to adapt to the global change. A recent anal-
ysis outlines that even pressures that do not kill any eels can have
impact s on eel populations by penalizing some ecotypes more than
others (Mateo, Lambert, Tétard, & Drouineau, 2017).
9 | OTHER IMPLICATIONS FOR EEL
MANAGEMENT AND RESEARCH
The eel de cline due to global c hange has several imp lications for man -
agement. First, global causes mean global solutions are warranted. By
global solutions, we do not mean that there should be a unique set of
management measures across all distribution areas, but rather coor-
dinated international management acting on each source of anthro-
pogenic pressure. This was proposed in the Quebec declaration of
concern (Dekker & Casselman, 2014; Dekker et al., 2003) that called
for immediate action and coordination at all scales. Although some
progress has been made since the first declaration, there is clear
need to improve management coordination among regional, national
and international authorities. Dekker (2016) pointed out the difficul-
ties in the implementation of the Eel Management Plan in Europe.
International coordination has not yet started for the American eel
(Castonguay & Durif, 2016; Jacoby et al., 2014; MacGregor et al.,
2008, 2009). The East Asia Eel Resource Consor tium does not yet
have any official support (Jacoby & Gollock, 2014b), and the first
attempt at international coordination took place in 2014 bet ween
South Korea, China, Taiwan and Japan with an agreement on the
amount of glass eel that c an be used for aquaculture.
Second, although it is difficult to disentangle the relative effects
of various anthropogenic pressures implicated in the decline, it is
DROUINE AU Et Al.
important to develop tools and methods to monitor and quantify
their effects in the future. Eels grow in very small and almost inde-
pendent units corresponding to river catchments (Dekker, 2000b)
with specific anthropogenic pressures, within which eels have dif-
ferent life history traits. Therefore, it is dif ficult to assess the stock
and extrapolate the overall effect of anthropogenic pressure at the
population scale, from observations collected at the river catch-
ment scale (Dekker, 2000a). However, the improvement in data
quality and the recent development of a generic model that can
be used at a larger geographic scale is a first step. For example,
the GEREM model provides estimates of glass eel recruitment that
can be used to assess glass eel fishery exploitation rates (Bornarel
et al., 2018; Drouineau, Beaulaton, et al., 2016). The models EDA
(Briand, Beaulaton, Chapon, Drouineau, & Lambert, 2015) or SMEP
(Aprahamian, Walker, Williams, Bark, & Knights, 20 07) can be used
to assess the abundance of yellow eels in river catchment s. These
can then be coupled with other models to assess spawner escape-
ment and the effect of different anthropogenic pressures such as
hydropower production or fisheries (Jouanin et al., 2012). Stock as-
sessment models have also been proposed to support management
(Bevacqua & De Leo, 20 06; Bevacqua, Melià, Gatto, & De Leo, 2015;
Dekker, 200 0a; Oeberst & Fladung, 2012). However, few tools are
currently available to assess the impact of contaminants on eel
populations. In a similar manner, there is a lack of tools to quan-
tify the effect of lost habitats on population dynamics, although
some methodologies are available which can quantify the amount
of habitat lost due to fragmentation. Although it is not possible to
quantify the historical effects of anthropogenic pressures, quan-
tifying and predicting pressures in the future would provide valu-
able information to prioritize management actions. Quantification
FIGURE10 Adaptation mechanisms to environmental heterogeneity as proposed in Mateo, Lambert, Tétard, Castonguay, et al. (2017),
Mateo, Lambert, Tétard, and Drouineau (2017), Gagnaire et al. (2012), Côté et al. (2014), Drouineau et al. (2014) and Boivin et al. (2015). A
red arrow stands for “unfavourable,” a green arrow stands for “favourable.” A blue arrow stands for a relationship which is either favourable”
or “unfavourable” depending on situations. There is a double arrow between genotypes and “settlement in upstream habitats” because it
represents “spatially varying selection” and “genetic- dependent habitat selection.” Regarding phenotypes, female is considered as opposite
to male and “settlement in upstream habitats” as opposite to “set tlement in downstream habitats”
DROUINEAU Et Al .
is even more important as (a) it is not possible to mitigate some of
the pressures affecting eels (parasitism, climate change), so it is
necessary to compensate their effects with mitigation measures
on the other pressures (fisher y, fragmentation, contamination); (b)
management practices cannot mitigate anthropogenic pressures at
similar temporal scales: Reduction in fishing efforts is recent but is
thought to operate quickly, whereas effor t to mitigate contamina-
tion or fragmentation is older but is much more complex and longer
Of course, temperate eels are not the only species endangered
by global change and most diadromous fishes have undergone
severe declines (Limburg & Waldman, 2009; McDowall, 1999;
Mota et al., 2015). The effects of fragmentation (Hax ton & Cano,
2016; Larinier, 2001; Limburg & Waldman, 2009), global warm-
ing (Elliott & Elliott, 2010; Friedland, 1998; Friedland, Hansen,
Dunkley, & MacLean, 2000; Jonsson & Jonsson, 2009; Lassalle,
Béguer, Beaulaton, & Rochard, 2008; Lassalle et al., 2009; Rougier
et al., 2014), fisheries and pollution (Limburg & Waldman, 20 09;
McDowall, 1999) have been documented for most of these spe-
cies. More generally, most migratory animals regardless of taxa
have undergone similar declines (Berger, Young, & Berger, 2008;
Sanderson, Donald, Pain, Burfield, & van Bommel, 2006; Wilcove
& Wikelski, 2008) raising the question of sustainability of migra-
tory tactics in the face of global change. In this context, why should
eels be considered as a symbol of the effec t of global change?
Because the original life cycle of eels make them vulnerable to
all five components of global change, and the cumulated impacts
of those five components have outpaced the adaptive c apacities
of these species acquired through million years of evolution. The
rate of change during the Great Acceleration in the second half of
the twentieth century was too fast for the adaptive capacity of
the eel, especially because the five components of global change
acted simultaneously. It explains how a species that was consid-
ered a vermin species in French salmonid rivers until the 1980s
has become critically endangered in only 25 years, after millions
of years of existence.
A total number of eels passing Moses- Saunders Hydroelectric Dam
are monitored and reported by Ontario Power Generation and
the New York Power Authority since 2006 and were provided by
Alastair Mathers. Hilaire Drouineau and Kazuki Yokouchi were par-
tially supported by the fund for international exchange and collabo-
ration of the Japan Fisheries Research and Education Agency. We
thank Miran Aprahamian for his help and the improvements he made
to the manuscript. We also would like to thank the editor and two
anonymous referees for their suggestions and comments.
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