consequences of the introduction of exotic and translocated species … · 2018. 8. 31. · exotic...

RESEARCHPAPER

Consequences of the introduction ofexotic and translocated species andfuture extirpations on the functionaldiversity of freshwater fish assemblagesShin-ichiro S. Matsuzaki1*, Takehiro Sasaki2 and Munemitsu Akasaka3

1Center for Environmental Biology and

Ecosystem Studies, National Institute for

Environmental Studies, 16-2 Onogawa,

Tsukuba-shi, Ibaraki 305-8506, Japan,2Graduate School of Frontier Sciences, The

University of Tokyo, 5-1-5 Kashiwanoha,

Kashiwa, Chiba 277-8653, Japan, 3Department

of Ecoregion Science, Tokyo University of

Agriculture and Technology, 3-5-8 Saiwai-cho,

Fuchu-shi, Tokyo 183-8509, Japan

ABSTRACT

Aim To explore the effects of the introduction of exotic and translocated speciesand possible future extirpation of native species on the functional diversity (FD) offreshwater fish assemblages.

Location Japanese archipelago.

Methods We examined spatio-temporal changes in species richness, FD, func-tional richness (the number of trait-based functional groups), and the functionalgroup composition between historical and current fish assemblages for 27 eco-regions, and compared the relative effects of the introduction of exotic and trans-located species on FD. We also used a null model approach to determine theassembly patterns and the extent of functional redundancy. Finally, we determinedthe effect of the loss of endangered species on FD by comparing the observed losseswith simulated random loss.

Results Through the introductions of non-native species, the species richness, FDand functional richness of the fish assemblages increased 2.4-, 1.6- and 2.1-fold,respectively. The functional group composition also changed largely through theadditions of new functional groups. Exotic species had a significantly greater effectsize than translocated species, but there were no differences in the overall net effectsof exotic and translocated species. Null modelling approaches showed that theobserved FD was higher than expected by chance (i.e. trait divergent) in bothhistorical and current assemblages. There was also low functional redundancy. Inour simulation, FD decreased in proportion to the loss of species, independent ofwhether the species were endangered.

Main conclusions We demonstrated that both exotic and translocated speciesmay change FD and functional group composition, which might have dramaticconsequences for ecosystem processes. We suggest that the future extirpation ofeven a few native species can cause a substantial loss of FD. Our findings emphasizethe need to improve conservation strategies based on species richness and conser-vation status, and to incorporate translocated species into targets of the manage-ment of non-native species.

KeywordsEcosystem functioning, fish assemblages, functional traits, Japan, limitingsimilarity, scenario, species introductions, species loss, stocking, traitdivergence.

*Correspondence: Shin-ichiro S. Matsuzaki,Center for Environmental Biology andEcosystem Studies, National Institute forEnvironmental Studies, 16-2 Onogawa,Tsukuba-shi, Ibaraki 305-8506, Japan.E-mail: [email protected]

bs_bs_banner

Global Ecology and Biogeography, (Global Ecol. Biogeogr.) (2013) 22, 1071–1082

© 2013 John Wiley & Sons Ltd DOI: 10.1111/geb.12067http://wileyonlinelibrary.com/journal/geb 1071

INTRODUCTION

Human activities have largely modified the composition of

biotas worldwide through the introduction of non-native

species and native extirpation (Winter et al., 2009; Villeger et al.,

2011), creating ‘novel assemblages’ that differ in composition

and/or function from past systems (Hobbs et al., 2006).

Although species richness is decreasing on a global scale, species

richness at the local and regional scales increases when the

number of invading non-native species greatly exceeds the

number of native species becoming extinct (Sax et al., 2002;

Winter et al., 2009; Baiser & Lockwood, 2011). Despite recent

advances in this area, the impact of invasion on existing biota

has often been underestimated because invasions of ‘translo-

cated species’ (i.e. species introduced within their native biogeo-

graphical zone in localities where they did not historically

occur) are typically excluded from the quantifications (however,

see Leprieur et al., 2008b). Compared with ‘exotic species’ (i.e.

species originating in another biogeographical zone), the impact

of the introduction of translocated species on ecosystem proc-

esses and functions has been much less widely considered (Guo

& Ricklefs, 2010; Carey et al., 2012).

In macroecology and invasive ecology, there is a growing need

to understand and predict the effects of species introductions on

ecosystem processes on a large scale (Simberloff & Rejmánek,

2011; Strayer, 2012). Taxonomic diversity indices implicitly

assume that all species contribute equally to ecosystem func-

tioning; these indices are silent on the functional and pheno-

typic differences between species. Functional diversity (hereafter

referred to as FD), which is the value and range of functional

traits of the organisms present in a given ecosystem, is consid-

ered to be an important determinant of ecosystem processes

(Diaz & Cabido, 2001; Petchey & Gaston, 2002a; Cadotte et al.,

2011). Because FD explains the effects of organisms on ecosys-

tems and provides a mechanistic link between organisms and

ecosystems, a number of studies have focused exclusively on the

consequences of species loss on FD. Nonetheless, the conse-

quences of species addition on FD and composition have rarely

been documented either empirically or theoretically.

The introduction of non-native freshwater fishes is a signifi-

cant component of global environmental change (Leprieur

et al., 2008a,b; Marr et al., 2010). This phenomenon has also

been observed for Japanese freshwater fish assemblages. The

Japanese archipelago exhibits a high endemism of freshwater

fishes; in this geographical area, speciation occurred in isolation

after colonization from the eastern Eurasian mainland because

of the restriction of their dispersal by physical barriers (oceans

and river divides) (Watanabe, 2012). In the last 100 years,

however, the biogeography of Japanese freshwater fishes has

been reshuffled dramatically as a result of human-mediated

introductions of both exotic and translocated species, while

regional extirpation of one fish species occurred in only 2 of 27

regions (Watanabe, 2012). At least 17 exotic species, including

largemouth bass (Micropterus salmoides), which is often inten-

tionally introduced for recreational or aquaculture purposes,

have been successfully established in Japan. Furthermore, inten-

tional and accidental translocations of many native fish species

have widely occurred. A major route for the introduction of

these native species is the contamination of stock enhancement

of commercially important fish, such as ayu (Plecoglossus altiv-

elis) and Japanese crucian carp (Carassius cuvieri) (Katano &

Matsuzaki, 2012). Ayu, which is the most important freshwater

fish to Japanese inland fisheries, has been vigorously stocked

into many rivers since 1913 using artificially reared seed fish and

young fish from Lake Biwa, an ancient lake and the largest lake

in Japan (Watanabe, 2012).

Although the addition of non-native freshwater fish species is

expected to increase overall community FD, changes in FD

depend on community assembly processes (Petchey et al., 2007;

Mendez et al., 2012). Freshwater fish display a remarkable range

and variance in morphological, behavioural, physiological,

trophic and life history traits (Mims et al., 2010). Because several

previous studies have suggested that niche partitioning is the

dominant process that shapes freshwater fish communities and

that coexisting species are functionally dissimilar (trait diver-

gence) (Mason et al., 2008a,b), we predicted that the overall FD

of freshwater fish communities would be high originally and

that the introduction of non-native species would further

increase the overall FD. Whether the overall FD increases pro-

portionally to species addition may depend on the functional

uniqueness of exotic and translocated species relative to native

species. However, even if the introduced species are not func-

tionally unique, increasing their richness would substantially

increase the overall FD.

Although only a few regional extirpations of native fish

species have been reported (Watanabe, 2012), future extirpation

of endangered fish species at the regional scale is considered to

be a realistic scenario because extirpation of these species can

occur with a substantial delay following habitat loss or degrada-

tion (Kuussaari et al., 2009). Predicting the effects of the future

extirpation of endangered species on processes within ecosys-

tems is essential to identifying the conservation priorities of

local communities (Gross & Cardinale, 2005; Sundstrom et al.,

2012). Continuous measures of FD offer the opportunity to

simulate the effect of possible scenarios of species loss on FD

(Petchey & Gaston, 2002b). If endangered species are function-

ally unique, the extirpation of these endangered species may

cause a substantial loss of FD that differs from that expected

from a random loss of species. To determine realistic conserva-

tion goals for maintaining the FD of Japanese freshwater fish

fauna at a large scale, it is necessary to simultaneously consider

the effects of the introduction of exotic and translocated species

and the effects of future extirpation of native species on FD.

Based on historical (1910s, i.e. prior to extirpations and intro-

ductions) and current (2010s, i.e. after extirpations and intro-

ductions) comprehensive datasets of strictly Japanese freshwater

fish faunas for 27 eco-regions of Japan, we explored the effects of

introductions of exotic and translocated species, and future

extirpations of endangered species on the FD of fish fauna. From

a macroecological perspective, our study had the following three

primary objectives. First, we investigated spatio-temporal

changes in FD, and the richness and composition of trait-based

S.-S. Matsuzaki et al.

Global Ecology and Biogeography, 22, 1071–1082, © 2013 John Wiley & Sons Ltd1072

functional groups. We also compared the relative effects of

exotic and translocated species on FD. Second, using a null

modelling approach, we investigated community assembly pat-

terns (trait divergence or convergence) and assessed the extent of

functional redundancy for fish assemblages. Finally, we investi-

gated the effect of the loss of endangered species on FD by

comparing plausible loss of species with simulated random loss.

METHODS

Fish distribution data and the identification ofnon-native species

We used historical and current datasets of strictly freshwater fish

distributions across 27 eco-regions, divided based on straits and

primary watersheds (Watanabe, 2012) (Fig. 1). The term ‘his-

torical data’ refers to the pre-industrial distribution of Japanese

native fish species; the intentional and accidental translocation

of fish species, such as ayu, occurred primarily after 1913. The

term ‘current data’ refers to present fauna (2010s) consisting of

native and non-native fishes after introductions and extirpa-

tions. For the current data, we partly updated the dataset of

Watanabe (2012) as follows. First, the distribution data of Achei-

lognathus macropterus was added because its successful estab-

lishment in the Kanto region has been confirmed (Hagiwara,

2002). Second, subspecies of the Carassius auratus complex were

not considered.

Political boundaries are considered to be important in halting

biological invasions and developing conservation strategies. In

the particular case of Japan, which is an insular system, the

political boundary has a biogeographical meaning. Thus, we

distinguished three categories of species: natives, exotic (i.e.

species originating from countries outside Japan) and translo-

cated (i.e. species native to Japan introduced into drainages

where they did not historically occur; in other words, domesti-

cally translocated). A total of 93 native species, including 38

translocated species, and 17 exotic species were included in our

analysis (Table S1 in Supporting Information).

Fish traits and measurement of observed FD

To calculate the community FD of each region, we chose 16

traits reflecting resource uses and life history strategies exhibited

by freshwater fish (Table 1). The traits chosen here were consid-

ered to be key determinants of ecosystem functioning in

freshwater ecosystems. Trait information was collated from

Natsumeda et al. (2010), Kawanabe & Mizuno (1989), Miyadi

et al. (1996), Nakamura (1969), the Japanese Red Data Book 2nd

edition (Ministry of Environment of Japan, 2003) and the Fish-

Base database (http://www.fishbase.org). Trophic guild and

dietary component were divided into independent binary traits

because these characteristics are not mutually exclusive (Petchey

et al., 2007).

The species ¥ trait matrix was converted into a distance

matrix, which was clustered to produce a dendrogram describ-

ing the functional relationships between species (Fig. S1)

(Petchey & Gaston, 2002a). Gower distance was used in this

study because Gower distance is appropriate for mixed (cat-

egorical and continuous) traits. To check the degree of distor-

tion engendered by clustering algorithms, we calculated the

cophenetic correlation between the ultrametric distance matrix

and the initial dissimilarity distance matrix. We also used a more

suitable goodness-of-fit measure based on a matrix norm, the

two-norm measure, which is based on matrix norm (Mérigot

et al., 2010). Unweighted pair group method using arithmetic

averages clustering was used because it had the highest cophe-

netic correlation (0.81, Fig. S2) and the lowest two-norm value

(3.3) of the clustering methods tested (Ward method, 284.6;

single linkage, 16.8; complete linkage, 21.5; weighted pair group

centroid method, 16.9; unweighted pair group centroid method,

17.0).

We calculated four observed FD values for each region:

FDhistorical, the FD of the historical assemblage; FDcurrent, the FD of

25

2423

2221

18

19

17

16

15

9

14

10

11

1

2

3 7

6W

8

5W

13

412

20

6E 5E

Figure 1 Regional division of the main islands of Japan and EastAsia used in the present analysis. The division of Japan, which isthe same as in Watanabe (2012), was as follows: 1, south-easternKyushu; 2, south-western Kyushu; 3, north-western Kyushu;4, north-eastern Kyushu; 5W, south-western Shikoku; 5E,south-eastern Shikoku; 6W, north-western Shikoku; 6E,north-eastern Shikoku: 7, south-western Chugoku Region; 8,south-eastern Chugoku Region (eastern Sanyo); 9, northernChugoku Region (San’ in); 10, middle Kinki Region; 11, northernKinki Region; 12, southern Kinki Region; 13, Tokai Regionaround Ise Bay; 14, eastern Tokai Region; 15, western HokurikuRegion; 16, eastern Hokuriku Region; 17, Kanto Region; 18,Pacific side of Tohoku Region; 19, Japan Sea side of TohokuRegion; 20, southern south-western Hokkaido; 21, northernsouth-western Hokkaido; 22, southern middle Hokkaido; 23,northern middle Hokkaido; 24, south-eastern Hokkaido; and 25,north-eastern Hokkaido.

Introductions and future extirpations affect freshwater fish FD

Global Ecology and Biogeography, 22, 1071–1082, © 2013 John Wiley & Sons Ltd 1073

the current assemblage; FDexotic, the FD of the current assem-

blage assuming that only exotic species have been introduced;

and FDtranslocated, the FD of the current assemblage assuming that

only translocated species have been introduced. We explored

temporal changes in the FD of fish assemblages by comparing

FDhistorical and FDcurrent. These two FD values were spatially

mapped using the gis software ArcGIS 10.0 program (Environ-

mental Systems Research Institute, Redlands, CA, USA). To

compare the effects of exotic and translocated species on FD, we

calculated the net effect (total increase in FD by increase in

translocated or exotic species) and effect size (the degree of

change in FD). The net effects of exotic and translocated species

were calculated as the differences between FDexotic or translocated and

FDhistorical, respectively. The effect sizes of exotic and translocated

species were also measured as the log response ratio (LRR)

(Micheli & Halpern, 2005) as follows:

LRFD FD

R lnexoticexotic historical

the number of increased= −( )

exotic species

LRFD FD

R lntranslocatedtranslocated historical

the number = −( )

oof increased translocated species

Tests for differences in the net effect and effect size between

exotic and translocated species were performed using a paired

t-test.

Null model methods, assembly patterns andfunctional redundancy

We compared patterns of observed FD of the assemblages with

patterns of expected FD of random communities to investigate

assembly patterns in the historical and current assemblages, and

to determine the extent of ecological redundancy (Micheli &

Halpern, 2005; Petchey et al., 2007; Sasaki et al., 2009; Brown &

Milner, 2012). To calculate expected FD, we constructed the

random assemblages, controlling for the number of species and

assuming that each species from the regional pool had the same

chance of occurring. In this case, 999 independent randomiza-

tions were performed per region. FD was calculated for each of

the 999 random assemblages, and the mean was used as the

expected FD for comparison with the observed FD. We used a

standardized measure of comparison between observed and

expected, the standardized effect size index (SES), calculated as

(observed FD - expected FD)/standard deviation of expected

FD (Thompson et al., 2010; Mendez et al., 2012). The assem-

blages are considered to be assembled at random if SES is not

significantly different from zero. SES > 0 suggests the prevalence

of trait divergence patterns, while SES < 0 suggests trait conver-

gence. We subsequently used a one-tailed Wilcoxon test to

determine whether the mean values of SES were significantly

different from zero.

To assess the extent of functional redundancy for Japanese

freshwater fish assemblages, we first compared the observed

changes in FD with the observed change in species richness

using linear regression. Temporal changes in species richness

and FD were quantified as LRR between the species richness and

FD, respectively, of the historical and current fish fauna. This

approach provides information on intrinsic redundancy result-

ing from functional similarity among taxa. Assemblages con-

taining many similar taxa have high intrinsic redundancy, and

random taxonomic changes may have little effect on FD. Second,

we compared the observed change in FD with the change in FD

predicted at random (see above). The LRR was also used in this

analysis. This approach provides information on extrinsic

redundancy, which can occur when non-random compositional

changes are non-random with respect to functional traits

(Petchey et al., 2007). When extrinsic redundancy does not exist,

the loss of relatively unique taxa causes a relatively large decrease

in FD.

Table 1 Traits used to measure functional diversity with regard to resource use and life history strategy.

Trait Type Values or range Units/categories Classification

Maximum total body length Continuous 3.2–150 Centimetre Adult

Body shape Categorical 4 categories Compressed, fusiform, cylindrical, dorso-ventrally flattened Adult

Trophic guild Binary 5 categories Omnivore, herbivore, detritivore, invertivore, piscivore Adult

Dietary component Binary 6 categories Plant/algae/detritus, zoobenthos, plankton, aquatic and

terrestrial insect, fish/egg, amphibian/mammals/bird

During a fish’s lifetime

Diet breadth Continuous 1–6 Total number of major diet items consumed During a fish’s lifetime

Foraging period Binary 2 categories Diurnal, nocturnal Adult

Vertical position Categorical 3 categories Demersal, benthopelagic, pelagic Adult

Temperature Binary 2 categories Warm, cool Adult

Flow Categorical 3 categories Fast, moderate, slow Adult

Substrate Categorical 4 categories General, silt/mud, sand, rubble Adult

Maturation Continuous 0.1–6 Year Adult

Parental protection Binary 2 categories Care, no care Adult

Egg diameter Categorical 3 categories 0–1.5 mm, 1.5–3.0 mm, > 3.0 mm Egg

Longevity Categorical 4 categories < 1 year, 2–4 years, 5–9 years, > 10 years During a fish’s lifetime

Fecundity Categorical 3 categories 0–1000 eggs, 1000–100,000 eggs, > 100,000 eggs Adult

Spawning substrate Categorical 5 categories Vegetation, mussel, mineral substrate, pelagic, various Adult



Compositional changes in trait-basedfunctional groups

To explore temporal changes in the functional compositions of

fish assemblages, we examined compositional changes in func-

tional groups between historical and current fish fauna. We

classified species into functional groups using the functional

dendrogram of the current assemblage that included both native

and non-native species. This method is often used in multivari-

ate approaches to dividing species among functional groups

(Fukami et al., 2005). We identified 19 clusters as functional

groups (Fig. S1). The cut-off for the number of clusters was

determined based on biological interpretation (Table S2). The

number of functional groups was also used as a measure of

functional richness.

We used principal component analysis (PCA) to investigate

compositional changes in the functional group composition

between historical and current fish fauna. We first generated a

presence/absence matrix for each functional group for each

region in each of the historical and current time periods. Thus,

we defined the main direction of variation in the functional

group composition across 27 regions in two time periods.

Extirpation simulations

Besides the effects of introductions of non-native species on FD,

we simulated the effects of the extirpation of endangered species

on FD based on future realistic scenarios. We predicted the

functional consequences of species extirpation following the

simulation method of Petchey & Gaston (2002b). Two extirpa-

tion scenarios that differ in the order of species loss were simu-

lated: (1) all species, both native and non-native, in a given

community become extinct in random order (randomly

ordered extirpation); and (2) only species within an endangered

species group in a given community become extinct in random

order (selected ordered extirpation). We targeted 65 endangered

species listed as category I (endangered) in national or prefec-

tural red lists and red data books. The two simulations were

performed using both historical and current assemblages for

each region. Species loss was simulated with 999 replicates, and

the number of simulations for each region was determined by

the number of endangered species (range 1–24 species among

regions, mean 6.1). We did not conduct simulations for regions

20 and 21, each of which included only one endangered species.

Finally, we assessed the statistical significance of the differences

Historical Current

Fu

ction

al d

ivers

ity

0

2

4

6

8

Historical Current

Sp

ecie

s r

ich

ness

0

10

20

30

40

Historical Current

Fu

nction

al rich

ness

0

2

4

6

8

10

12

14

*

*

*

(d) (g)

Historical Current

)i()f(

(d)

(e)

(f)

(g)

(h)

(i)

(a)

(b)

(c)

Species richness1–20

21–40

41–60

> 60

Species richness1–20

21–40

41–60

> 60

Functional richness1–5

6–10

11–15

16–20

Functional richness1–5

6–10

11–15

16–20

Functional diversity0–2.50

2.51–5.00

5.01–7.50

> 7.51

Functional diversity0–2.50

2.51–5.00

5.01–7.50

> 7.51

Figure 2 Changes in species richness (a,d, g), functional diversity (FD; b, e, h)and functional richness (number oftrait-based functional groups; c, f, i) ofJapanese freshwater fish assemblagesbetween historical and current timeperiods. Bars represent one standarderror. Asterisks denote significance level:*P < 0.05; n.s., not significant (pairedt-test). Note that regions 20 and 21,mapped in (e), are shown in whitebecause the presence of only one speciesdoes not permit the calculation of FD.



in FD values between randomly and selected ordered simula-

tions using one-sided Monte Carlo tests with a = 0.05.

RESULTS

Spatio-temporal changes in FD and functionalgroup composition

The species richness of current fish assemblages has increased

2.4-fold, on average, compared with their historical assemblages

(Fig. 2a, d, g; paired t-test P < 0.01) because of the introduction

of non-native species. With increases in non-native species rich-

ness, the observed FD of the current assemblages was on average

1.6-fold higher than that of the historical assemblages (Fig. 2b, e,

h; paired t-test P < 0.01). This result indicates that the current

assemblages occupy a greater functional trait space than histori-

cal assemblages. Similarly, the functional richness of the current

assemblages was on average 2.1-fold higher than that of the

historical assemblages (Fig. 2c, f, i; paired t-test P < 0.01),

indicating that new functional groups were added to the

assemblages.

The net effects of translocated and exotic species

(FDtranslocated – FDhistorical vs. FDexotic – FDhistorical) were not signifi-

cantly different (Fig. 3a, paired t-test P = 0.14). However, the

effect size of exotic species (LRRexotic) was significantly greater

than that of translocated species (Fig. 3b, paired t-test P < 0.01),

indicating that the introduction of exotic species produce a

greater change in FD, on average, than the introduction of trans-

located species.

PCA clearly showed differences in functional group compo-

sition of historical and current fish fauna (Fig. 4). The first two

principal components accounted for 61.1% of the total varia-

tion in functional group compositions, but only the first axis

(PC1) explained 43.9% of the sampled variance. PC1 separated

the current assemblages from the historical assemblages. From

historical to current time periods, the scores moved to the right

of the plot (positive loadings). Functional groups 1, 2, 6 and 10

were most strongly negatively correlated with PC1 (with factor

loadings of 0.36, 0.35, 0.36 and 0.33, respectively). Functional

groups 1, 6 and 10 were composed of two or three exotic

species, while functional group 2 included two exotic species

and one translocated species (Fig. S1 and Table S2). In the five

regions (21, 22, 23, 24 and 25) of Hokkaido Island, composi-

tional changes in the functional groups were relatively small

(Fig. 4).

Trait distribution patterns and functionalredundancy patterns

In both the historical and current time periods, a strong and

significant linear relationship between species richness and FD

was evident (paired t-test P < 0.01). The observed FDs of the fish

assemblages were higher than the expected FDs in almost all

regions in both historical and current time periods. The SES

values were significantly greater than zero (Fig. 5a, b; one-tailed

Wilcoxon test P < 0.01), suggesting trait divergence. The 95%

confidence intervals of the SESs of the historical and current

time periods also overlapped.

Strong and significant relationships were evident between

observed changes in observed FD, species richness and changes

in expected FD (Fig. 6a, b; paired t-test P < 0.01). The slope of

the relationship between changes in observed FD and increases

in species richness was 0.82 (� 0.04 SE), which was close to 1

(95% confidence interval for slope = 0.77–0.87). This result

indicates low intrinsic redundancy. The slope of the relationship

between changes in observed FD and changes in expected FD

was 1.15 (� 0.06 SE) and overlapped 1 (95% confidence interval

for slope = 1.10–1.20). This result indicates a lack of extrinsic

redundancy.

LRR

do

me

stica

lly tra

nslo

ca

ted

or

exo

tic

–2.5

–2.0

–1.5

–1.0

–0.5

0.0

ExoticTranslocated

Net

eff

ect

(tota

l in

cre

ase in

FD

)

0.0

0.5

1.0

1.5

2.0

(a)

(b)

n.s.

Translocated Exotic

*Figure 3 Effects of translocated and exotic species on functionaldiversity (FD). (a) The mean net effects (� SE) of translocatedand exotic species on FD (total increase in FD; FDtranslocated or

exotic – FDhistorical, respectively); (b) the effect sizes of translocatedand exotic species (the degree of change in FD; LRRtranslocated andLRRexotic). The mean effect size (� SE) was quantified using thelog response ratio. Asterisks denote significance level: *P < 0.05;n.s., not significant (paired t-test).



Extirpation simulations

In all regions, both randomly ordered and selected ordered

species extirpations in both the historical and current assem-

blages caused a rapid loss of FD (Fig. 7). Even initial extirpations

gave rise to proportional losses of FD, and FD decreased pro-

portionally with subsequent loss of species. This finding reflects

the absence of functional redundancy. In almost all regions in

both the historical and current time periods, the losses of FD in

selected ordered extirpations of endangered species were similar

to the losses that occurred in randomly ordered species extirpa-

tions. This finding suggests that endangered native species may

not be functionally unique and that species extirpation order

may have no significant effects on the pattern of FD loss.

However, in the current assemblages of regions 6W and 7, FD

loss was significantly lower when selected ordered extirpation

occurred than when randomly ordered species extirpation

occurred (a = 0.05), when more than one and seven endangered

species, respectively, became extinct.

DISCUSSION

To our knowledge, this study is the first to simultaneously dem-

onstrate the effects of introductions of exotic and translocated

species, and future extirpation of native species on FD on a

national scale. Our analysis showed that the introduction of

non-native fish species has resulted in striking increases in the

FD and functional richness of Japanese strictly freshwater fish

assemblages (Fig. 2) and has profoundly changed their func-

tional group compositions (Fig. 4). Although the introduction

of exotic species had a significantly greater effect size than the

Figure 4 Principal component analysis(PCA) biplots and correspondingloadings for the trait-based functionalgroup composition of Japanesefreshwater fish assemblages at 27eco-regions during historical and currenttime periods. The numbers adjacent tothe circles represent the individualregions. The codes on the arrows arefunctional groups, and functional groupscorrespond to Fig. S1.

Historical Current

0.0

0.5

1.0

1.5

2.0

2.5

SE

S (

mean ±

95%

CI)

Figure 5 Patterns of deviation of the standard effect size (SES)from zero in the historical (a) and current (b) fish assemblagesacross 27 regions. Mean values of SES and 95% confidenceintervals are represented; the grey solid circles show the SESs foreach eco-region.



introduction of translocated species, there was no difference in

the overall net effect between exotic and translocated species

(Fig. 3). These changes in FD and functional group composition

might have large consequences for ecosystem processes and

functions. Furthermore, we demonstrated that the future extir-

pation of native species, independent of whether they are

endangered, may cause a substantial loss of FD in Japanese fish

assemblages (Fig. 7), likely because of the absence of functional

redundancy (Fig. 6).

Effects of exotic and translocated species on FD

The introduction of exotic species had significantly greater

effect size than the introduction of translocated species

(Fig. 3b), suggesting that exotic species are relatively function-

ally unique compared with native species. Exotic species origi-

nating from the Nearctic, such as largemouth bass and bluegill,

were likely to have longer branch lengths in the functional den-

drogram compared with exotic species from the same realm as

Japan (Fig. S1). However, even though the ‘exotic species’

included many species from the same realm as Japan, we found

a clear difference in the effect size between exotic and translo-

cated species. In contrast, there was no difference in the overall

net effect produced by translocated and exotic species (Fig. 3a).

This difference could be caused by two factors. First, there was

no significant difference in the richness of the translocated and

exotic species (P = 0.44). The high richness of the translocated

species may contribute to their net effect. Second, some species

endemic to Lake Biwa, a representative ancient lake in East Asia,

were introduced widely, most likely through stocking of ayu.

Because those endemic species showed relatively functional

uniqueness compared with other native species, their introduc-

tion may have had substantial effects on FD. Our findings

suggest that the effects of exotic and translocated species on

FD may change depending on their functional uniqueness and

richness.

The addition of new functional groups based on traits can

often significantly impact ecosystem processes, functions and

stability (Diaz & Cabido, 2001). In the majority of regions in the

present study, functional group composition changed markedly

with the introduction of new functional groups (Fig. 4). Func-

tional groups 1 and 6, which have high PC1 scores, are charac-

terized by large-bodied exotic fishes (Table S2). Blanchet et al.

(2010) suggested that humans tend to introduce large-bodied

species for angling, aquaculture and fisheries. Changes in func-

tional group composition may also be associated with changes

in dietary traits; for example, functional groups 1, 2 and 6, which

have high PC1 scores, are piscivores and species with a wide diet

breadth (Table S2). In Japan, where there are few native pisciv-

ores (Azuma & Motomura, 1998), an increase in the piscivore

occurrence could seriously affect ecosystem processes through a

trophic cascade. Indeed, at the local scale, the introduction of

non-native piscivore largemouth bass caused the extirpation of

native fish species and broad changes in native freshwater com-

munities (Maezono & Miyashita, 2003).

Potential assembly rules

The SESs of the fish assemblages across Japan were greater than

zero, regardless of the time period (Fig. 5); co-occurring species

were more dissimilar in their functional traits than expected.

Although we cannot definitively identify the specific mecha-

nisms underlying these patterns, niche partitioning may act as

one of the mechanisms structuring the historical assemblages.

Mason et al. (2008b) suggests that niche partitioning is an

important mechanism for species coexistence in freshwater fish

communities, although it should be kept in mind that their

study was conducted in Europe and on a lake scale. Although

Schleuter et al. (2012) reported that the SESs of European fresh-

water fish communities vary depending on environmental con-

LRR (Change in species richness)

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4

LRR

(C

ha

nge

in o

bse

rve

d F

D)

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

Y = 0.82 X - 0.03

(r2 = 0.93, P < 0.01)

(r2 = 0.96, P < 0.01)

LRR (change in expected FD)

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4

LRR

(C

ha

nge

in o

bse

rve

d F

D)

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

Y = 1.15 X - 0.07

(a)

(b)

Figure 6 Temporal changes in species richness and functionaldiversity (FD) between historical and current time periods. (a)Relationship between observed changes in species richness andFD. (b) Relationship between expected and observed changes inFD. Solid lines denote the linear regression model relationship;dashed lines denote a 1:1 relationship. LRR, log response ratio.



ditions, almost all SES values of the historical Japanese fish

assemblages studied here exceeded zero, most likely because our

observed environmental range is limited. It is important to note

that the relative effects of niche partitioning would operate pre-

dominantly at a local scale (Mendez et al., 2012). Further study

is required to examine the scale dependency of the assembly

processes.

The assembly process of the current assemblages must be

interpreted carefully because the species evolved in their envi-

ronment for a sufficiently long period for the selection of func-

tional attributes to occur. The stocking of ayu began in 1913 and

many exotic species were successfully established in the 1970s

and 1980s. Because it is likely that current fish assemblages have

not reached equilibrium in terms of community trait structure,

it may be expected that niche partitioning is still structuring

current fish assemblages. Therefore, the consequences of the

introductions of exotic and translocated species for trait assem-

bly should be examined in the longer term.

Future extirpation and functional redundancy

Our simulations demonstrate that future extirpations have a

potentially large effect on FD (Fig. 7). The response of FD to

extirpation depends on whether each species is functionally

unique or whether some degree of functional redundancy exists

(Blackburn et al., 2005). The relationship between species rich-

ness and FD emphasizes that there is relatively little evidence of

functional redundancy among fish species (Fig. 6). Two factors

appear to contribute to the lack of any observed redundancy.

First, the fish assemblages studied had low overlap in their func-

tional traits among component species (Fig. 5), suggesting low

intrinsic redundancy (Micheli & Halpern, 2005). Second, there

was no evidence of extrinsic redundancy (Fig. 6). In fact, there

was evidence of sensitivity; i.e. the relationship between the

observed and the expected change in FD was nearly 1 (Petchey

et al., 2007). Thus, even in species-rich fish assemblages, the

extirpation of a few species can result in a significant loss of FD.

Although some previous studies have examined the func-

tional consequences of species extirpation on FD (Petchey &

Gaston, 2002b; Gross & Cardinale, 2005; Batalha et al., 2010),

few studies have evaluated whether effective conservation of

endangered species is essential for the maintenance of ecosystem

processes and functions (but see Sundstrom et al., 2012).

Endangered species are generally considered to be both evolu-

tionarily distinct and ecologically unique (Redding et al., 2010).

However, in this study, endangered species were not functionally

unique compared with common species, suggesting that it is

important to note the effects of depletion of common species

and endangered species on FD (Gaston & Fuller, 2008).

Although Sundstrom et al. (2012) reported that almost com-

plete functionality (functional group composition) within com-

munities is conserved despite the extirpation of endangered

species (i.e. the redundancy of functional groups), our findings

may differ from theirs because there was no evidence of func-

tional redundancy in the Japanese freshwater fish community.

Caveats and future issues

At least two caveats related to the work presented here provide a

focus for future research. First, our conclusion that the invasion

of non-native species increases FD stems from work conducted

at one spatial scale; in the eco-regional scale we employed,

regional extirpation of native fish species rarely occurred

(Watanabe, 2012). Because stocking of commercially important

fish species has been conducted widely throughout Japan, the

effects of introduction of non-native species on FD would not

(a) Random extirepation from historical assemblages

Number of extirepated species

0 5 10 15 20 25

FD

0

2

4

6

8

10

(c) Selected extirepation from historical assemblages


0 5 10 15 20 25

FD

0

2

4

6

8

10

1

2

3

4

5W

5E

6W

6E

7

8

9

10

11

12

13

14

15

16

17

18

19

22

23

24

25

(b) Random extirepation from current assemblages


0 5 10 15 20 25

FD

0

2

4

6

8

10

(d) Selected extirepation from current assemblages


0 5 10 15 20 25

FD

0

2

4

6

8

10Figure 7 Effects of (a, b) randomlyordered and (c, d) selected orderedextirpations on functional diversity (FD)in (a, c) historical and(b, d) currentJapanese freshwater fish assemblages. Theselected ordered extirpations simulatedthe extirpation of only endangeredspecies. Colours correspond withnumbered regional divisions provided inFig. 1.



be expected to differ greatly at other scales. An increase in the

richness of exotic and translocated species in many Japanese

lakes and rivers has been reported. However, the extirpation of

native species occurs more frequently on the local scale (lakes)

than the regional scale used here (S.S. Matsuzaki, unpublished

data). If our results were applied at the local scale, the effects of

extirpation on FD would be underestimated. Considering the

results of extirpation simulation, the reduction of native species

could affect FD and functional group composition more signifi-

cantly at local scales.

Another important caveat of the present work is that we relied

only on species presence–absence data and did not consider

relative abundance data. Devictor et al. (2010) suggested that

variations in species abundance are generally ignored when cal-

culating FD. Even relatively small proportional increases in the

abundance of non-native species can have a dramatic impact on

ecosystem structure and function. Although it is generally diffi-

cult to obtain precise abundance data at the regional scale, the

use of abundance-weighted FD on smaller scales may be highly

relevant in cases in which dominant, non-native species affect

important ecosystem processes.

Implications for conservation

Understanding and predicting FD patterns is essential for

improving current conservation strategies based on species

richness (Devictor et al., 2010; Cadotte et al., 2011; Strecker

et al., 2011). The Japanese Ministry of the Environment

enacted a new ‘Invasive Alien Species (IAS) Act’ in June 2006;

under this law, only exotic species are defined as IAS, and such

species can no longer be legally imported, introduced, trans-

ferred, bred or released in the field. However, the effectiveness

of this law is compromised because it does not encompass

translocated species. In addition to ensuring that contamina-

tion of other native species through annual stocking of com-

mercially important fish such as ayu is avoided, these

introductions should be stringently regulated or reduced to the

lowest level possible. Furthermore, the ecological impacts of

translocated species may be insufficiently understood by the

public and decision makers.

Both past and current conservation strategies almost exclu-

sively focus on endangered species in their attempts to protect

rarity and endemism (Devictor et al., 2010). Our simulation

results emphasize the importance of the considering conserva-

tion of full biota from the perspective of FD. In addition, our

redundancy analysis suggests that conserving a large proportion

of the functional traits of species requires conserving a large

proportion of all native species. Although efficient and proactive

conservation planning should take these complexities into

account, it is logistically and financially infeasible to conserve all

native species. O’Gorman et al. (2011) demonstrated that the

loss of highly unique, weakly interacting species may lead to

dramatic and unpredictable levels of ecosystem functions; thus,

the identification of such species could be essential. To maintain

biodiversity, ecosystem processes and functions, we believe that

regions that are minimally invaded should be protected and that

the management of native species that are highly imperilled and

functionally distinct should be prioritized.

ACKNOWLEDGEMENTS

We are most grateful to Takaharu Natsumeda for supplying trait

data, and to Katsutoshi Watanabe, David Currie, Janne Soininen

and the three anonymous referees for their important com-

ments and suggestions. This study was supported by the Envi-

ronment Research and Technology Development Fund (S9) of

the Ministry of the Environment, Japan, with additional support

from Tohoku University’s Global COE program (no. J03) from

the Ministry of Education, Culture, Sports, Science and Tech-

nology of Japan.

REFERENCES

Azuma, M. & Motomura, Y. (1998) Feeding habits of large-

mouth bass in a non-native environment: the case of a small

lake with bluegill in Japan. Environmental Biology of Fishes, 52,

379–389.

Baiser, B. & Lockwood, J.L. (2011) The relationship between

functional and taxonomic homogenization. Global Ecology

and Biogeography, 20, 134–144.

Batalha, M.A., Cianciaruso, M.V. & Motta-Junior, J.C.

(2010) Consequences of simulated loss of open cerrado

areas to bird functional diversity. Natureza & Conservacao, 8,

34–40.

Blackburn, T.M., Petchey, O.L., Cassey, P. & Gaston, K.J. (2005)

Functional diversity of mammalian predators and extinction

in island birds. Ecology, 86, 2916–2923.

Blanchet, S., Grenouillet, G., Beauchard, O., Tedesco, P.A., Lep-

rieur, F., Dürr, H.H., Busson, F., Oberdorff, T. & Brosse, S.

(2010) Non-native species disrupt the worldwide patterns of

freshwater fish body size: implications for Bergmann’s rule.

Ecology Letters, 13, 421–431.

Brown, L.E. & Milner, A.M. (2012) Rapid loss of glacial ice

reveals stream community assembly processes. Global Change

Biology, 18, 2195–2204.

Cadotte, M.W., Carscadden, K. & Mirotchnick, N. (2011)

Beyond species: functional diversity and the maintenance of

ecological processes and services. Journal of Applied Ecology,

48, 1079–1087.

Carey, M.P., Sanderson, B.L., Barnas, K.A. & Olden, J.D. (2012)

Native invaders – challenges for science, management, policy,

and society. Frontiers in Ecology and Environments, 10, 373–

381.

Devictor, V., Mouillot, D., Meynard, C., Jiguet, F., Thuiller, W. &

Mouquet, N. (2010) Spatial mismatch and congruence

between taxonomic, phylogenetic and functional diversity:

the need for integrative conservation strategies in a changing

world. Ecology Letters, 13, 1030–1040.

Diaz, S. & Cabido, M. (2001) Vive la difference: plant functional

diversity matters to ecosystem processes. Trends in Ecology and

Evolution, 16, 646–655.



Fukami, T., Bezemer, T.M., Mortimer, S.R. & van der Putten,

W.H. (2005) Species divergence and trait convergence in

experimental plant community assembly. Ecology Letters, 8,

1283–1290.

Gaston, K.J. & Fuller, R.A. (2008) Commonness, population

depletion and conservation biology. Trends in Ecology and

Evolution, 23, 14–19.

Gross, K. & Cardinale, B.J. (2005) The functional consequences

of random vs. ordered species extinctions. Ecology Letters, 8,

409–418.

Guo, Q. & Ricklefs, R.E. (2010) Domesitc exotics and the per-

ception of invasibility. Diversity and Distributions, 16, 1034–

1039.

Hagiwara, T. (2002) Establishment of an introduced bitterling,

Acheilognathus macropterus, in Lake Kasumigaura, Japan.

Botejako, Osaka, 6, 19–22.

Hobbs, R.J., Arico, S., Aronson, J., Baron, J.S., Bridgewater, P.,

Cramer, V.A., Epstein, P.R., Ewel, J.J., Klink, C.A., Lugo, A.E.,

Norton, D., Ojima, D., Richardson, D.M., Sanderson, E.W.,

Valladares, F., Vila, M., Zamora, R. & Zobel, M. (2006) Novel

ecosystems: theoretical and management aspects of the new

ecological world order. Global Ecology and Biogeography, 15,

1–7.

Katano, O. & Matsuzaki, S.S. (2012) Biodiversity of freshwater

fishes in Japan in relation to inland fisheries. The biodiversity

observation network in Asia-Pacific region: towards further

development of monitoring (ed. by S. Nakano, T. Yahara and T.

Nakashizuka), pp. 431–444. Springer, Tokyo.

Kawanabe, H. & Mizuno, N. (1989) Freshwater fishes of Japan.

Yama-Kei Publishers Co. Ltd, Tokyo, Japan (in Japanese).

Kuussaari, M., Bommarco, R., Heikkinen, R.K., Helm, A.,

Krauss, J., Lindborg, R., Öckinger, E., Pärtel, M., Pino, J.,

Roda, F., Stefanescu, C., Teder, T., Zobel, M. & Steffan-

Dewenter, I. (2009) Extinction debt: a challenge for

biodiversity conservation. Trends in Ecology and Evolution, 24,

564–571.

Leprieur, F., Beauchard, O., Blanchet, S., Oberdorff, T. & Brosse,

S. (2008a) Fish invasions in the world’s river systems: when

natural processes are blurred by human activities. PLoS

Biology, 6, 404–410.

Leprieur, F., Beauchard, O., Hugueny, B., Grenouillet, G. &

Brosse, S. (2008b) Null model of biotic homogenization: a test

with the European freshwater fish fauna. Diversity and Distri-

butions, 14, 291–300.

Maezono, Y. & Miyashita, T. (2003) Community-level impacts

induced by introduced largemouth bass and bluegill in farm

ponds in Japan. Biological Conservation, 109, 111–121.

Marr, S.M., Marchetti, M.P., Olden, J.D., García-Berthou, E.,

Morgan, D.L., Arismendi, I., Day, J.A., Griffiths, C.L. &

Skelton, P.H. (2010) Freshwater fish introductions in

mediterranean-climate regions: are there commonalities in

the conservation problem? Diversity and Distributions, 16,

606–619.

Mason, N.W.H., Lanoiselée, C., Mouillot, D., Wilson, J.B. &

Argillier, C. (2008a) Does niche overlap control relative abun-

dance in French lacustrine fish communities? A new method

incorporating functional traits. Journal of Animal Ecology, 77,

661–669.

Mason, N.W.H., Irz, P., Lanoiselée, C., Mouillot, D. & Argillier,

C. (2008b) Evidence that niche specialization explains

species-energy relationships in lake fish communities. Journal

of Animal Ecology, 77, 285–296.

Mendez, V., Gill, J.A., Burton, N.H.K., Austin, G.E., Petchey, O.L.

& Davies, R.G. (2012) Functional diversity across space and

time: trends in wader communities on British estuaries. Diver-

sity and Distributions, 18, 356–365.

Mérigot, B., Durbec, J. & Gaertner, J. (2010) On goodness-of-fit

measure for dendrogram-based analyses. Ecology, 91, 1850–

1859.

Micheli, F. & Halpern, B.S. (2005) Low functional redundancy in

coastal marine assemblages. Ecology Letters, 8, 391–400.

Mims, M.C., Olden, J.D., Shattuck, Z.R. & Poff, N.L. (2010) Life

history trait diversity of native freshwater fishes in North

America. Ecology of Freshwater Fish, 19, 390–400.

Ministry of Environment of Japan (2003) Threatened wildlife of

Japan, red data book, brackish and fresh water fishes, 2nd edn.

Ministry of Environment, Tokyo (in Japanese).

Miyadi, D., Kawanabe, H. & Mizuno, N. (1996) Colored illustra-

tion of the freshwater fishes of Japan. Hoikusha Publishing Co.,

Ltd., Osaka, Japan (in Japanese).

Nakamura, M. (1969) Nippon no koi-ka gyorui (Cyprinid fishes of

Japan). Research Institute of Natural Resources, Tokyo (in

Japanese).

Natsumeda, T., Tsuruta, T. & Iguchi, K. (2010) An evaluation of

the ecological features of endangered freshwater fish com-

monly distributed in Japan. Nippon Suisan Gakkaishi, 76, 169–

184. (in Japanese with English abstract).

O’Gorman, E.J., Yearsley, J.M., Crowe, T.P., Emmerson, M.C.,

Jacob, U. & Petchey, O.L. (2011) Loss of functionally unique

species may gradually undermine ecosystems. Proceedings of

the Royal Society B: Biological Sciences, 278, 1886–1893.

Petchey, O.L. & Gaston, K.J. (2002a) Functional diversity (FD),

species richness and community composition. Ecology Letters,

5, 402–411.

Petchey, O.L. & Gaston, K.J. (2002b) Extinction and the loss of

functional diversity. Proceedings of the Royal Society B: Biologi-

cal Sciences, 269, 1721–1727.

Petchey, O.L., Evans, K.L., Fishburn, I.S. & Gaston, K.J. (2007)

Low functional diversity and no redundancy in British avian

assemblages. Journal of Animal Ecology, 76, 977–985.

Redding, D.W., DeWolff, C.V. & Mooers, A.O. (2010) Evolution-

ary distinctiveness, threat status, and ecological oddity in pri-

mates. Conservation Biology, 24, 1052–1058.

Sasaki, T., Okubo, S., Okayasu, T., Jamsran, U., Ohkuro, T. &

Takeuchi, K. (2009) Two-phase functional redundancy in

plant communities along a grazing gradient in Mongolian

rangelands. Ecology, 90, 2598–2608.

Sax, D.F., Gaines, S.D. & Brown, J.H. (2002) Species invasions

exceed extinctions on islands worldwide: a comparative study

of plants and birds. The American Naturalist, 160, 766–783.

Schleuter, D., Daufresne, M., Veslot, J., Mason, N.W.H.,

Lanoiselée, C., Brosse, S., Beauchard, O. & Argillier, C. (2012)



Geographic isolation and climate govern the functional diver-

sity of native fish communities in European drainage basins.

Global Ecology and Biogeography, 21, 1083–1095.

Simberloff, D. & Rejmánek, M. (2011) Encyclopedia of biological

invasions. University of California Press Books, Berkeley.

Strayer, D.L. (2012) Eight questions about invasions and ecosys-

tem functioning. Ecology Letters, 15, 1199–1210.

Strecker, A.L., Olden, J.D., Whittier, J.B. & Paukert, C.P. (2011)

Defining conservation priorities for freshwater fishes accord-

ing to taxonomic, functional, and phylogenetic diversity. Eco-

logical Applications, 21, 3002–3013.

Sundstrom, S.M., Allen, C.R. & Barichievy, C. (2012) Species,

functional groups, and thresholds in ecological resilience.

Conservation Biology, 26, 305–314.

Thompson, K., Petchey, O.L., Askew, A.P., Dunnett, N.P., Beck-

erman, A.P. & Willis, A.J. (2010) Little evidence for limiting

similarity in a long-term study of a roadside plant commu-

nity. Journal of Ecology, 98, 480–487.

Villeger, S., Blanchet, S., Beauchard, O., Oberdorff, T. & Brosse,

S. (2011) Homogenization patterns of the world’s freshwater

fish faunas. Proceedings of the National Academy of Sciences

USA, 108, 18003–18008.

Watanabe, K. (2012) Faunal structure of Japanese freshwater

fishes and its artificial disturbance. Environmental Biology of

Fishes, 94, 533–547.

Winter, M., Schweiger, O., Klotz, S., Nentwig, W., Andriopoulos,

P., Arianoutsou, M., Basnou, C., Delipetrou, P., Didžiulis, V.,

Hejda, M., Hulme, P.E., Lambdon, P.W., Pergl, J., Pyšek, P.,

Roy, D.B. & Kühn, I. (2009) Plant extinctions and introduc-

tions lead to phylogenetic and taxonomic homogenization of

the European flora. Proceedings of the National Academy of

Sciences USA, 106, 21721–21725.

SUPPORTING INFORMATION

Additional supporting information may be found in the online

version of this article at the publisher’s web-site.

Figure S1 Functional relationships among 110 freshwater fish

species in Japan including native and exotic species. The

numbers correspond to fish ID in Table S1 (black in bold: exotic

species; grey in bold: translocated species; grey: native species).

The vertical line indicates the partitioning level defining 17

functional groups (see also text).

Figure S2 Relationship between the cophenetic (or ultrametric)

distance matrix derived from the dendrogram and the original

distance matrix based on the Gower distance.

Table S1 Freshwater fish species evaluated in the study, includ-

ing status, biogeographical realm, and historical and current

distributions.

Table S2 Summary of functional groups constructed based on

the dendrogram produced by UPGMA of the distance matrix

calculated from the functional traits of species.

BIOSKETCHES

Shin-ichiro S. Matsuzaki is a researcher at the

National Institute of Environmental Studies. He is

interested in the conservation of freshwater biodiversity

and the intersection of ecological theory with this

applied goal, with a taxonomic emphasis on freshwater

fishes.

Takehiro Sasaki is an assistant professor at Tokyo

University. He is interested in the consequences of

realistic loss of biodiversity on ecosystem functioning

and services.

Munemitsu Akasaka is a senior associate professor

at Tokyo University of Agriculture and Technology. He

is interested in the mechanisms and consequences of

biological invasion and conservation planning at the

regional and national scales, mainly focusing on

terrestrial and freshwater plants.

Editor: Janne Soininen



SUPPORTING INFORMATION 1

2

Figure S1 Functional relationships among 110 freshwater fish species in Japan including 3

native and exotic species. The numbers correspond to fish ID in Table S1 (black in bold: 4

exotic species; gray in bold: translocated species, gray: native species). The vertical line 5

indicates the partitioning level defining 17 functional groups (see also text). 6

7

Figure S2 Rrelationship between the cophenetic (or ultrametric) distance matrix derived 8

from the dendrogram and the original distance matrix based on the Gower distance. 9

10

Table S1 Freshwater fish species evaluated in the study, including status, biogeographic 11

realm, and historical and current distributions. 12

13

Table S2 Summary of functional groups constructed based on the dendrogram produced by 14

UPGMA of the distance matrix calculated from the functional traits of species. 15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

30

31

32

33

34

35

36

37

38

39

40

41

42

43

Figure S1 Matsuzaki et al. 44

103104106

21105

8184828388929386858799

108109

525370545571777675747273807879

100102101107

95949697394021

19272841424533323435222324505162696367666465684344464849374791151416171812

511

110437

13968

109858565761596029303189902025263638

0.0 0.1 0.2 0.3 0.4

Height

FG1

FG2

FG3

FG4

FG5

FG6

FG7

FG8

FG9FG10

FG11

FG12

FG13

FG14

FG15

FG16

FG17

FG18

FG19

45

46

47

48

49

50

51

52

53

54

55

56

57

58

59

60

61

62

63

64

65

Figure S2 Matsuzaki et al.66

0.0 0.1 0.2 0.3 0.4 0.5 0.6

0.0

0.1

0.2

0.3

0.4

The dissimilarity distance matrix based

The

coph

enet

ic d

ista

nce

mat

rix

Cophenetic correlation index = 0.81

Table S1 Freshwater fish species evaluated in the study, Status, biogeographic realm, and historical and current distributions. 67 68

ID Family Species Status Biogeographic realm

Historical distributions in Japan (see the region number in Fig.1)

No. of region

Historical Current

1 Cyprinidae Carassius cuvieri Native/Domestically translocated Palearctic 10 1 22

2 Cyprinidae Carassius auratus complex

Native/Domestically translocated Palearctic 1, 2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9,

10, 11, 12, 13, 14, 15, 16, 17, 18, 19 21 27

3 Cyprinidae Tanakia tanago Native Palearctic 17 1 1

4 Cyprinidae Tanakia lanceolata Native Palearctic 2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9, 10, 11, 12, 13, 14, 15, 16, 17, 18, 19 20 20

5 Cyprinidae Tanakia limbata Native/Domestically translocated Palearctic 2, 3, 4, 6W, 6E, 7, 8, 9, 10, 11, 12,

13 12 15

6 Cyprinidae Acheilognathus melanogaster Native Palearctic 17, 18 2 2

7 Cyprinidae Acheilognathus tabira tabira

Native/Domestically translocated Palearctic 6E, 8, 10, 13 4 7

8 Cyprinidae Acheilognathus tabira erythropterus Native Palearctic 17, 18 2 2

9 Cyprinidae Acheilognathus tabira tohokuensis Native Palearctic 16, 19 2 2

10 Cyprinidae Acheilognathus tabira jordani Native Palearctic 9, 11, 15 3 3

11 Cyprinidae Acheilognathus tabira nakamurae Native Palearctic 2, 3 2 2

12 Cyprinidae Acheilognathus rhombeus

Native/Domestically translocated Palearctic 2, 3, 8, 10, 11, 12 6 12

13 Cyprinidae Acheilognathus cyanostigma

Native/Domestically translocated Palearctic 8, 10, 11, 12, 13 5 13

14 Cyprinidae Acheilognathus longipinnis Native Palearctic 10, 13, 15 3 3

15 Cyprinidae Acheilognathus typus Native/Domestically translocated Palearctic 16, 17, 18, 19 4 5

16 Cyprinidae Rhodeus ocellatus kurumeus Native Palearctic 2, 3, 4, 6E, 7, 8, 10 7 7

17 Cyprinidae Rhodeus atremius atremius

Native/Domestically translocated Palearctic 2, 3, 4 3 4

18 Cyprinidae Rhodeus atremius suigensis Native Palearctic 8 1 1

19 Cyprinidae Ischikauia steenackeri Native/Domestically translocated Palearctic 10 1 12

20 Cyprinidae Hemigrammocypris rasborella

Native/Domestically translocated Palearctic 2, 3, 4, 6E, 8, 10, 13, 14 7 8

21 Cyprinidae Opsariichthys uncirostris Native/Domestically translocated Palearctic 10, 11 2 21

22 Cyprinidae Zacco platypus Native/Domestically translocated Palearctic 2, 3, 4, 6W, 6E, 7, 8, 9, 10, 11, 12,

13, 14, 15, 16, 17 16 21

23 Cyprinidae Zacco temminckii Native/Domestically translocated Palearctic 1, 2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9,

10, 11, 12, 13, 14, 15 17 19

24 Cyprinidae Zacco sieboldii Native/Domestically translocated Palearctic 2, 3, 4, 6E, 7, 8, 10, 11, 13, 15 10 12

25 Cyprinidae Aphyocypris chinensis Native Palearctic 2, 3 2 2

26 Cyprinidae Rhynchocypris percnurus sachalinensis Native Palearctic 22, 23, 24, 25 4 4

27 Cyprinidae Rhynchocypris lagowskii Native Palearctic 8, 10, 11, 12, 13, 14, 15, 16, 17, 18, 19 11 11

28 Cyprinidae Rhynchocypris oxycephalus

Native/Domestically translocated Palearctic 1, 2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9,

10, 11, 12, 13, 14, 15, 17 18 19

29 Cyprinidae Pseudorasbora parva Native/Domestically translocated Palearctic 1, 2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9,

10, 11, 12, 13, 14, 15, 16, 17 19 26

30 Cyprinidae Pseudorasbora pumila pumila Native Palearctic 16, 17, 18, 19 4 3

31 Cyprinidae Pseudorasbora pumila subsp.

Native/Domestically translocated Palearctic 13 1 2

32 Cyprinidae Sarcocheilichthys biwaensis Native Palearctic 10 1 1

33 Cyprinidae Sarcocheilichthys variegatus variegatus

Native/Domestically translocated Palearctic 2, 3, 7, 8, 9, 10, 11, 12, 13 9 11

34 Cyprinidae Sarcocheilichthys variegatus microoculus


35 Cyprinidae Pungtungia herzi Native/Domestically translocated Palearctic 2, 3, 4, 5E, 6E, 7, 8, 9, 10, 11, 12, 13 12 13

36 Cyprinidae Gnathopogon elongatus Native/Domestically translocated Palearctic 4, 5W, 6W, 6E, 7, 8, 9, 10, 11, 12,

13, 14, 15, 16, 17 15 20

37 Cyprinidae Gnathopogon caerulescens


38 Cyprinidae Gnathopogon suwae Native Palearctic 14 1 0

39 Cyprinidae Biwia zezera Native/Domestically translocated Palearctic 2, 3, 8, 10, 13 5 13

40 Cyprinidae Biwia sp. Native Palearctic 10 1 1

41 Cyprinidae Pseudogobio esocinus Native Palearctic 1, 2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9, 10, 11, 12, 13, 14, 15, 16, 17, 18, 19 21 21

42 Cyprinidae Abbottina rivularis Native/Domestically translocated Palearctic 2, 3, 8, 10, 11, 12, 13 7 10

43 Cyprinidae Hemibarbus longirostris Native/Domestically translocated Palearctic 6E, 7, 8, 9, 10, 11, 12, 13 8 10

44 Cyprinidae Hemibarbus labeo Native/Domestically translocated Palearctic 6E, 7, 8, 9, 10, 11, 12, 13 8 10

45 Cyprinidae Hemibarbus barbus Native/Domestically translocated Palearctic 2, 3, 6W, 7, 10, 14, 15, 16, 17, 18,

19 11 14

46 Cyprinidae Squalidus gracilis gracilis

Native/Domestically translocated Palearctic 2, 3, 4, 6W, 6E, 7, 8, 9, 10, 11, 12,

13 12 16

47 Cyprinidae Squalidus japonica japonica Native Palearctic 10, 13 2 2

48 Cyprinidae Squalidus chankaensis biwae


49 Cyprinidae Squalidus chankaensis subsp.

Native/Domestically translocated Palearctic 6E, 7, 8, 10, 12, 13 6 9

50 Cobitidae Leptobotia curta Native Palearctic 8, 10 2 2

51 Cobitidae Misgurnus anguillicaudatus

Native/Domestically translocated Palearctic

1, 2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9, 10, 11, 12, 13, 14, 15, 16, 17, 18, 19, 20, 21

23 27

52 Cobitidae Niwaella - 7 -elicate (Group G) Native Palearctic 10, 11, 13 3 3

53 Cobitidae Niwaella - 7 -elicate (Group S) Native Palearctic 15 1 1

54 Cobitidae Cobitis takatsuensis Native Palearctic 3, 4, 7, 9 4 4

55 Cobitidae Cobitis shikokuensis Native Palearctic 5W, 6W 2 2

56 Cobitidae Cobitis biwae (Eastern) Native Palearctic 17, 18, 19 3 3

57 Cobitidae Cobitis biwae (Kochi) Native Palearctic 5W, 5E 2 2

58 Cobitidae Cobitis biwae (Western) Native Palearctic 4, 6W, 6E, 7, 8, 9, 10, 11, 12, 13, 14, 15, 16 13 13

59 Cobitidae Cobitis sp. Y86 Native Palearctic 2 1 1

60 Cobitidae Cobitis sp. Y90 Native Palearctic 3 1 1

61 Cobitidae Cobitis sp. Y94 Native Palearctic 4, 7 2 2

62 Cobitidae Cobitis sp. 3 (Middle) Native Palearctic 3, 4, 6W, 6E, 7, 8, 9, 10, 11, 12 10 10

63 Cobitidae Cobitis sp. 2 subsp. 1 (Small-Sanyo) Native Palearctic 8 1 1

64 Cobitidae Cobitis sp. 2 subsp. 2 (Small-Tokai) Native Palearctic 13, 14 2 2

65 Cobitidae Cobitis sp. 2 subsp. 3 (Small-San’in) Native Palearctic 9 1 1

66 Cobitidae Cobitis sp. 2 subsp. 4 (Small-Kyushu) Native Palearctic 2 1 1

67 Cobitidae Cobitis sp. 2 subsp. 5 (Small-L.Biwa) Native Palearctic 10 1 1

68 Cobitidae Cobitis sp. 2 subsp. 6 (Small-R.Yodo) Native Palearctic 10 1 1

69 Cobitidae Cobitis sp. 1 (Large) Native Palearctic 10 1 1

70 Cobitidae Noemacheilus barbatulus Native/Domestic Palearctic 22, 23, 24, 25 4 7

toni ally translocated

71 Cobitidae Lefua nikkonis Native/Domestically translocated Palearctic 22, 23, 24, 25 4 10

72 Cobitidae Lefua echigonia (Hokuriku) Native Palearctic 16, 19 2 2

73 Cobitidae Lefua echigonia (Tohoku) Native Palearctic 18 1 1

74 Cobitidae Lefua echigonia (Tokai) Native Palearctic 13, 14 2 2

75 Cobitidae Lefua echigonia (Kinki) Native Palearctic 10, 11, 15 3 3

76 Cobitidae Lefua echigonia (N-Kanto) Native Palearctic 17, 18 2 2

77 Cobitidae Lefua echigonia (S-Kanto) Native Palearctic 17 1 1

78 Cobitidae Lefua sp. (Kinki-Shikoku) Native Palearctic 6E, 12 2 2

79 Cobitidae Lefua sp. (Sanyo) Native Palearctic 8, 9, 10, 11 4 4

80 Cobitidae Lefua sp. (Tokai) Native Palearctic 13 1 1

81 Bagridae Pseudobagrus nudiceps Native/Domestically translocated Palearctic 4, 6W, 6E, 7, 8, 9, 10, 11, 12 9 18

82 Bagridae Pseudobagrus tokiensis Native Palearctic 16, 17, 18, 19 4 4

83 Bagridae Pseudobagrus ichikawai Native Palearctic 13 1 1

84 Bagridae Pseudobagrus aurantiacus Native Palearctic 1, 2, 3 3 3

85 Siluridae Silurus lithophilus Native Palearctic 10 1 1

86 Siluridae Silurus biwaensis Native Palearctic 10 1 1

87 Siluridae Silurus asotus Native/Domestically translocated Palearctic 1, 2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9,

10, 11, 12, 13, 14, 15, 16, 17, 18, 19 21 23

88 Amblycipitidae Liobagrus reini Native Palearctic

2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9, 10, 11, 12, 13, 14, 15, 16, 17, 18, 19, 20, 21

20 20

89 Adrianichth Oryzias latipes subsp. Native/Domestic Palearctic 11, 15, 16, 19 4 5

yidae ally translocated

90 Adrianichthyidae Oryzias latipes latipes Native Palearctic 1, 2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9,

10, 11, 12, 13, 14, 17, 18, 19, 20, 21 18 18

91 Sinipercidae Coreoperca kawamebari Native/Domestically translocated Palearctic 2, 3, 4, 6E, 7, 8, 9, 10, 11 9 11

92 Odontobutiidae Odontobutis obscura Native/Domestic

ally translocated Palearctic 1, 2, 3, 4, 5W, 5E, 6W, 6E, 7, 8, 9, 10, 11, 12, 13, 15 16 18

93 Odontobutiidae Odontobutis hikimius Native Palearctic 9 1 1

94 Cyprinidae Ctenopharyngodon idellus Exotic Palearctic/Ori

ental - 0 11

95 Cyprinidae Mylopharyngodon piceus Exotic Palearctic/Oriental - 0 3

96 Cyprinidae Aristichthys nobilis Exotic Palearctic/Oriental - 0 4

97 Cyprinidae Hypophthalmichthys molitrix Exotic Palearctic/Ori

ental - 0 7

98 Cyprinidae Rhodeus ocellatus ocellatus Exotic Palearctic/Ori

ental - 0 22

99 Ictaluridae Ictalurus punctatus Exotic Nearctic - 0 4

100 Synbranchidae Monopterus albus Exotic Palearctic/Ori

ental - 0 7

101 Poeciliidae Gambusia affinis Exotic Nearctic - 0 16

102 Poeciliidae Poecilia reticulata Exotic Nearctic - 0 8

103 Centrarchidae Micropterus salmoides Exotic Nearctic - 0 23

104 Centrarchidae Micropterus dolomieu Exotic Nearctic - 0 10

105 Centrarchidae Lepomis macrochirus Exotic Nearctic - 0 22

106 Cichlidae Oreochromis spp. Exotic Afrotropical - 0 11

107 Belontiidae Macropodus chinensis Exotic Palearctic/Oriental - 0 4

108 Channidae Channa maculata Exotic Palearctic/Oriental - 0 6

109 Channidae Channa argus Exotic Palearctic/Oriental - 0 21

110 Cyprinidae Acheilognathus macropterus Exotic Palearctic/Ori

ental - 0 1

Table S2 Summary of functional groups constructed based on the dendrogram produced by UPGMA of the distance matrix calculated from 69 the functional traits of species. 70 71

Functional group

Number of species

Group description Exotic Native

(Domestically translocated)

1 2 0 (0) Large body size, piscivores, egg protection 2 2 1 (1) Compressed body shape, piscivores, wide diet breadth

3 0 7 (2) Dorso-ventrally flattened body shape, nocturnal, prefer rubble habitats, spawn eggs on mineral substrates

4 0 3 (1) Large body size, Dorso-ventrally flattened body shape, piscivores, nocturnal

5 1 0 (0) Dorso-ventrally flattened body shape, large body size, nocturnal, long longevity

6 2 0 (0) Large body size, piscivores, wide diet breadth, nocturnal 7 0 2 (0) Prefer fast current velocity, herbivores/detritivores

8 0 13 (2) Cylindrical body shape, invertivores, prefer cold, moderate current velocity and sand environments

9 1 0 (0) Wide diet breadth, long longevity, spawn in pelagic habitats 10 3 0 (0) Prefer pelagic habitats and spawn in pelagic habitats

11 4 0 (0) Large body size, herbivores or invertivores, spawn in pelagic habitats, high fecundity and long longevity

12 0 2 (1) Dermersal, short longevity, prefer sand environments 13 0 3 (3) Large body size, herbivores or omnivores, spawn eggs on vegetation, high

fecundity, moderate longevity

14 0 12 (9) Fusiform, omonivores, feed on terresitial insects, benthopelagic, prefer rubble and sand habitats

15 0 1 (0) Demersal with compressed body shape, prefer rubble habitats, spawn in various substrates

16 0 16 (7) Benthopelagc, invertivores, spawn eggs on various substrates, prefer slow current velocity and mud/silt habitats

17 0 1 (1) Invertivores, prefer rubble habitats and slow current velocity enviroments

18 2 18 (6) Small body size, compressed body shape, spawn eggs on mussels, herbivores/detritivores,

19 0 16 (5) Small body size, benthopelagic, omnivores, spawn in meneral substrates or vegitations.

72

73

74

75

76

77

78

consequences of the introduction of exotic and translocated species … · 2018. 8. 31. · exotic...

Documents