Effects of multi-year droughts on fish assemblages of seasonally drying Mediterranean streams

Effects of multi-year droughts on fish assemblages ofseasonally drying Mediterranean streams

M. FILOMENA MAGALHAES* , PEDRO BEJA †, ‡, ISAAC J . SCHLOSSER § AND

MARIA J. COLLARES-PEREIRA*

*Universidade de Lisboa, Faculdade de Ciencias, Centro de Biologia Animal, Campo Grande, Lisboa, Portugal†ERENA, Rua Robalo Gouveia, Lisboa, Portugal‡CIBIO, Centro de Investigacao em Biodiversidade e Recursos Geneticos, Campus Agrario de Vairao, Universidade do Porto,

Vairao, Portugal§Department of Biology, University Station, Grand Forks, ND, U.S.A.

SUMMARY

1. This study analysed changes occurring in Mediterranean stream fish assemblages over a

sequence of dry years followed by a generally wet period (1991–98). Variations in

assemblage attributes were quantified at the basin and stream reach scales, and related to

variables reflecting the occurrence of unusually dry or wet conditions.

2. Assemblage variability increased along with the resolution of analysis, with little

changes in species richness, composition and rank abundances, but significant variation in

individual species abundances. Fluctuations in relative abundances were significantly

affected by variables reflecting the severity of summer droughts and the occurrence of

rainy springs. These patterns were evident at the basin scale, while variability at individual

stream reaches tended to be higher and less related to rainfall patterns.

3. At least three response guilds to rainfall variation could be identified: two of the four

abundant and widespread species (chub and loach) declined following dry years, whereas

the two other core species (nase and eel) declined after rainy spring; one scarce native

species (stickleback) increased in dry years.

4. Except at the two most upstream reaches, the assemblages tended to recover quickly to

previous configuration after the changes occurring during the sequence of dry years.

5. Temporal variability of local assemblages was concordant among reaches but did not

follow any consistent spatial pattern. Instead, spatial patterns in assemblage attributes

changed over time in response to environmental variability, with a tendency for a

disruption of upstream–downstream gradients following dry years.

6. Results supported the view that present-day droughts cause relatively small and

transient changes to Mediterranean stream fish assemblages. However, longer and more

severe droughts expected under altered future climates, may result in declines or local

extinctions of the most sensitive species and their potential replacement by more resistant

species. Changing drought regimes thus need to be duly considered in the development of

conservation strategies for Mediterranean stream fish.

Keywords: climate change, community dynamics, conservation, disturbance, resilience

Introduction

In the Mediterranean basin, several studies have

revealed general trends towards reduced precipita-

tion totals and increased inter-annual variability

(Schonwiese & Rapp, 1997; IPCC, 2001; Millan, Estrela

Correspondence: Maria Filomena Magalhaes, Centro de Biologia

Ambiental, Departamento de Biologia Animal, Faculdade de

Ciencias de Lisboa, Campo Grande, 1749-016 Lisboa, Portugal.

E-mail: [email protected]

Freshwater Biology (2007) 52, 1494–1510 doi:10.1111/j.1365-2427.2007.01781.x

1494 � 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd

& Miro, 2005), which are consistent with scenarios of

future climate changes (Gibelin & Deque, 2003; Pal,

Giorgi & Bi, 2004; Santos & Miranda, 2006). As a

consequence, it is expected that the magnitude,

frequency and spatial area of droughts over the

region will increase in the future (Jones et al., 1996).

These likely increases in the number and severity of

extreme drought events, exacerbated by growing

demands for water by agricultural, industrial and

tourist activities (e.g. Laraus, 2004), will probably

have extensive impacts on Mediterranean freshwater

ecosystems.

Predicting the ecological consequences of changing

drought regimes in Mediterranean streams is challen-

ging, because these ecosystems are intrinsically char-

acterised by predictable seasonal events of summer

drying over an annual cycle, which vary markedly in

timing, intensity and frequency on a multi-annual

scale in association with rainfall patterns (Gasith &

Resh, 1999). Concurrently, floods during the wet

season are also common, further contributing for a

high natural variability of flow conditions (Gasith &

Resh, 1999). In these circumstances, it might be

expected that organisms inhabiting these historically

harsh and highly variable systems would be resistant

to droughts, because consistent selective or filtering

pressure on the type of species attributes appropriate

for local persistence might have been at work for long

periods of evolutionary time (Poff, 1997). However,

many Mediterranean aquatic species seem to be living

already at the edge of their tolerance limits because of

anthropogenic impacts, such as damming, water

abstraction, pollution and the spread of exotic species

(e.g. Cowx & Collares-Pereira, 2000; Gil-Sanchez &

Alba-Tercedor, 2006; Smith & Darwall, 2006), and may

be vulnerable to further stressors like the increased

frequency of extreme drought events. Studies of

droughts carried out now may help overcoming these

uncertainties, thereby providing managers with in-

formation required for the conservation of Mediterra-

nean freshwater biodiversity under altered future

environments (Matthews & Marsh-Matthews, 2003).

Understanding the effects of extreme droughts is

particularly important in the case of Mediterranean

stream fish assemblages, which include a large of

number of steadily declining endemic species (e.g.

Oberdorff, Guegan & Hugueny, 1995; Cowx &

Collares-Pereira, 2000; Smith & Darwall, 2006).

Furthermore, studies carried out in both temperate

and dry regions have reported frequently the occur-

rence of significant drought effects, including changes

in fish assemblage richness and composition, in

species relative abundances, and in population age

and size structure (e.g. Grossman et al., 1998; Schlos-

ser et al., 2000; Grossman et al., 2006; review in

Matthews & Marsh-Matthews, 2003). Although dry-

ness is a normal constraint in the lives of Mediterra-

nean fish, there is some evidence that droughts may

also result in significant population fluctuations

(Bernardo et al., 2003; Magalhaes, Schlosser & Col-

lares-Pereira, 2003). At present, however, there is still

little information on the degree to which these

assemblages are altered following droughts (resis-

tance), and the degree to which they return to their

previous configuration after the end of drought

conditions (resilience). These responses will likely be

conditional upon drought magnitude, including the

water deficit in relation to years of average precipita-

tion and the duration of the dry spell. Moreover,

assemblages may be particularly sensitive to the

cumulative effects of droughts occurring over a

number of consecutive years (Matthews & Marsh-

Matthews, 2003). It is expected that the severity of

drought effects may vary along the longitudinal

stream gradient, as downstream reaches are generally

assumed to have more stable environmental condi-

tions along with increased species richness and

abundances (Matthews, 1998; Magalhaes, Batalha &

Collares-Pereira, 2002a).

This study examined these issues, using an 8-year

(1991–98) data set of fish samples collected in a small

Mediterranean basin in southern Portugal. During the

study, the region was affected by severe drought

conditions in the hydrological years (October–Sep-

tember) of 1991/1992, 1992/1993 and 1994/1995

(Paulo, Pereira & Matias, 2003). These dry years were

followed by a generally wet period, including a very

wet year in 1995/1996. Floods occurred in some wet

years, with a particularly severe flash flood in

November 1997 following a precipitation episode of

over 100 mm in about 7 h (100-years return period;

Rodrigues, Brandao & Alvares, 1998). In a previous

study it was shown that this environmental variability

was associated with marked population fluctuations

of two dominant cyprinid species (Magalhaes et al.,

2003). The present paper examined the degree to

which the overall assemblage structure (number,

identity and relative abundance of species) changed

Effects of droughts on Mediterranean stream fish 1495

� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 1494–1510

over the succession of dry years, and the degree to

which the previous configuration recovered during

the wet period. These patterns were assessed at the

basin scale and at individual reaches, testing whether

responses to droughts varied with spatial scale and

along the longitudinal stream gradient. In addition,

upstream–downstream patterns of assemblage com-

position and assemblage variability were assessed.

Results were then used to discuss the implications of

changing drought regimes on Mediterranean stream

fish.

Methods

Study area

The study area and fish assemblages have been

described in detail elsewhere (Beja, 1996; Magalhaes

et al., 2002b, 2003). Briefly, the study was conducted

in the basin (238 km2) of a small stream (Torgal;

28.0 km) discharging at sea level in the Mira estuary

(SW Portugal). The basin drains siliceous igneous

rocks, slates and greywackes. Human settlement is

sparse and open oak woodlands, eucalyptus forestry

and extensive agriculture dominates land cover.

Riparian galleries are generally well preserved and

the stream is largely free from urban pollution,

impoundments, angling and other recreational activ-

ities. Climate is Mediterranean, with annual rainfall

varying markedly from year to year (298–1120 mm;

Fig. 1); about 80% of the annual rain occurs in

October–March and virtually none in the hot, dry

months (July–August). Mean monthly temperature

ranges from 11 �C (December) to 24 �C (August). The

flow regime is highly dependent on rainfall patterns.

Headwaters flow only after heavy rains, while the

main stream and the largest tributaries typically dry

to isolated pools and aquifer-fed shallow runs in the

dry season. In dry years, there are no significant

floods, the drying period is extended, and surface

water is restricted to the deepest pools. In wet years

there may be major floods and flows may persist

through the dry season in downstream reaches. Fish

assemblages include a reduced pool of native species,

dominated by two endemic cyprinids: chub (Squalius

torgalensis Coelho, Bogutskaya, Rodrigues and Coll-

ares-Pereira) and nase (Chondrostoma almacai Coelho,

Mesquita and Collares-Pereira). Eel (Anguilla anguilla

L.) and loach (Cobitis paludica De Buen) are also

abundant, whereas barbel (Barbus sclateri Gunther)

and stickleback (Gasterosteus gymnurus Cuvier) occur

sparsely and at low numbers. Exotic species are

relatively rare and include gambusia (Gambusia hol-

brooki Girard), pumpkinseed (Lepomis gibbosus L.) and

largemouth bass (Micropterus salmoides Lacepede).

Fish populations in this basin are largely isolated, as

no freshwater species other than eel occur regularly in

the Mira estuary (Costa et al., 1994).

Fish sampling

Fish were sampled yearly from 1991 to 1998 at four

50-m reaches along the Torgal stream (T1–4) and two

50-m reaches in each of its two main tributaries:

Capelinha (C1–2) and Vale Ferro (F1–2) (Fig. 2).

Reaches were located in stream sections maintaining

either isolated pools or reduced water flows during

the dry season, and they were selected to be repre-

sentative of habitat diversity and for ease of access.

Upstream ephemeral reaches were not sampled, as

they generally lack fish (M.F. Magalhaes, unpublished

data). Each year, sampling was carried out during

approximately 2 weeks in April–July, after the rainy

period and before the summer dry-up, when flow

conditions were most stable and similar among

reaches. As a consequence, sampling was a few weeks

earlier in dry years than in wet years, so that flow

conditions were as comparable as possible across

years (Magalhaes et al., 2003).

Each reach was fished for 1 h, using a single

anode electro-fishing gear (350 V, 3–4A, DC). Fish

were identified to species, counted, measured for

total length (to the nearest mm), and returned to the

stream. Preliminary investigation showed that this

procedure yielded on average 90.7% of the species

(SD ¼ 11.9% n ¼ 21, range: 66.7–100%) recorded in

three consecutive electrofishing passes, and there

were high correlations between the first and the

overall sample in both species ranks (RS; Mean ¼0.98, SD ¼ 0.034, n ¼ 20, range ¼ 0.90–1.0) and relat-

ive abundances (R; Mean ¼ 0.97, SD ¼ 0.041, n ¼20, range ¼ 0.85–1.0) (M.F. Magalhaes & P. Beja,

unpublished data). Participation of the same field

crew throughout the study and training of assist-

ants ensured that methods were consistent among

years.

Fish abundances in a given year and stream reach

were indexed from the number of individuals of each

1496 M.F. Magalhaes et al.


species caught therein. Annual abundances at the

basin scale were then indexed by pooling the catch

data from all sampling reaches. Counts of individual

fish species were directly used in analysis, because

stream reach lengths and sampling methods were the

same across reaches. Densities (counts per unit area)

were not used, because temporal fluctuations in

density could have resulted from changes in fish

numbers, changes in habitat area, or both, thereby

making it impossible to assess the actual effects of

droughts on fish populations. In order to compensate

for potential biases associated with variation of

sampling dates across years, age-0 fish were identified

using length frequency distributions and excluded

from analysis.

Drought quantification

Drought conditions are notoriously difficult to quan-

tify, with most approaches reflecting the interaction

between the availability and need of water for human

uses, and not necessarily any relevant ecological

process (e.g. McMahon & Finlayson, 2003). To

circumvent this problem, this study used various

parameters derived from annual rainfall data to index

the shortage of water for aquatic organisms. Data

(c)

(b)

(a)

Fig. 1 Time series (1931–99) for the Torgal

basin (Relıquias station) of: (a) annual

rainfall (October–September) and cumu-

lative negative deviation from the median

(water deficit); (b) wet season rainfall

(October–April) and positive deviation

from the median (surplus); and (c) spring

rainfall (April–May) and positive devi-

ation from the median (surplus). The long-

term median and 25% quartile are provi-

ded for each rainfall variable, as well as

the temporal trend fitted using negative

exponentially-weighted least squares.



were obtained from the gauging station of Relıquias

(37�42¢N; 8�29¢W; 230 m elevation), where it was

available the longest time series (1931 – present) near

the study area. Rainfall was used instead of flow data,

because there was no stream gauging station in the

study area. Furthermore, the most severe drought

effects probably occurred during the hot and dry

summer period, when in many reaches the flow

ceased and fish was confined to isolated pool refugia

(Magalhaes et al., 2002b). Because of this, analyses

were based on hydrologic years from the beginning of

the wet season (October 1) to the end of the dry season

(September 30), instead of calendar years. Two meas-

ures of drought severity were the negative deviations

of annual rainfall from the corresponding long-term

median and 25% quartile, with positive values reset to

zero. Two additional parameters were computed as

the run-sums of the same negative deviations, with

positive values also reset to zero (Yevjevich, 1967).

This approach is based on the theory of runs (e.g.

Moye et al., 1988), and it has the advantage over

simple deficits of accounting for eventual cumulative,

inter-annual effects of low rainfall on the shortage of

water in a basin (Yevjevich, 1967). In the analyses,

these four variables were used against fish catches in

the following year, because fish surveys were carried

out before the dry season. As the occurrence of high

flows in some years could confound the interpretation

Years

Torgal basin

91 92 93 94 95 96 97 980

100

200

300

hsifforeb

muN

BassGambusiaEelSticklebackLoachBarbelChubNase

91 92 93 94 95 96 97 980

100

200

91 92 93 94 95 96 97 980

100

200

300

400

500

91 92 93 94 95 96 97 980

100

200

300

91 92 93 94 95 96 97 980

100

200

300

91 92 93 94 95 96 97 980

100

200

300

91 92 93 94 95 96 97 980

100

200

300

91 92 93 94 95 96 97 980

100

200

300

0 5 km

Reach (T1)

Reach (T2)

Reach (T3)

Reach (C2)

Reach (C1)

91 92 93 94 95 96 97 980

100

200Reach (F1)

Reach (F2)

Reach (T4)

Fig. 2 Changes in fish abundance between 1991 and 1998 in the Torgal watershed (SW Portugal), at the basin scale and at each of eight

50-m reaches.



of drought effects, simple rainfall surpluses were

computed for both the wet season (October–April)

and spring (April–May), considering both long-term

medians and 25% quartiles. Spring was considered on

its own, as it tends to be the primary fish-spawning

period (e.g. Crivelli & Britton, 1987; Oliva-Paterna,

Torralva & Fernandez-Delgado, 2002; Magalhaes

et al., 2003).

Data analyses

A set of quantitative approaches of increasing resolu-

tion was used to examine assemblage patterns and to

identify the factors influencing change over the

sequence of dry and wet years. Various methods

were selected to allow comparisons with previous

studies, to incorporate different information and to

encompass variability in a range of assemblage

attributes from species composition to relative abun-

dances (Grossman, Dowd & Crawford, 1990; Mat-

thews, 1998; Eby, Fagan & Minckley, 2003). Analysis

were conducted using assemblage attributes esti-

mated at both the basin and the reach scales, because

assemblage patterns occur at different spatial scales

and can vary with the scale of observation (Leibold

et al., 2004).

Persistence in assemblage composition was quanti-

fied using an index of species turnover rate (T)

between time periods, calculated as T ¼ (C + E)/

(S1 + S2), where C and E are the number of species

that colonised or were extirpated and S1 and S2 are

the number of species present in each period (Eby

et al., 2003). Constancy in species rank abundances

was tested using Kendall’s coefficient of concordance

(W; Siegel & Castellan, 1988). Rank concordances were

interpreted using their statistical significance and

magnitude, with high W-values (0.75–1.0) taken to

indicate high assemblage stability, and intermediate

(0.50–0.75) and low (0.0–0.50) values reflecting mod-

erate and low stability, respectively (Grossman et al.,

1990). Variability of individual species abundances

was quantified using the coefficient of variation

[CV ¼ (SD/mean) · 100]: CV £ 25% ¼ highly stable;

25% < CV £ 50% ¼ moderately stable; 50% < CV £75% ¼ moderately fluctuating; CV > 75% ¼ highly

fluctuating (Grossman et al., 1990).

A time lag regression analysis (Collins, 2000) was

used to examine whether the assemblages were

undergoing a long-term directional changes or a

directional change followed by a return to a previous

condition (cyclic pattern). The method involved cal-

culation of the Euclidean Distance (ED) between each

possible pair of annual species abundance samples.

Linear and quadratic regressions were then used to

relate ED values to the square root of the time lag

separating the samples. The square root transforma-

tion reduces the potential for few points at larger time

lags biasing the analysis. A significant linear regres-

sion indicates that the assemblage is undergoing a

directional change, whereas a significant quadratic

regression provides support to the presence of a drift-

and-recovery pattern.

Principal component analysis (PCA) and redund-

ancy analysis (RDA) were used to examine the

temporal trajectories of assemblage composition and

to evaluate whether assemblage changes were related

to rainfall deficits and surpluses. Linear ordination

methods were selected because preliminary detrend-

ed correspondence analyses showed turnovers <2SD,

which is the recommended criterion for choosing

linear versus uni-modal ordination models (terBraak,

1995). The analyses were carried out on correlation

matrices, to give similar weighting to all species.

Species data were transformed as log10 (X + 1) to

reduce the influence of peak catches in the abundance

of dominant species. The significance of each envi-

ronmental variable in determining assemblage chan-

ges were assessed in RDA conducted at the basin

scale and at each of the eight sampling reaches.

Statistical significance was determined by Monte

Carlo permutations tests (104 unrestricted permuta-

tions). Model building used the forward selection

procedure available in the CANOCOCANOCO 4 software

(terBraak & Smilauer, 1998), with significant

(P £ 0.05) explanatory variables added to the model

in the order of greatest additional contribution to the

total variation explained.

Spatial patterns in assemblage composition and

variability were examined using linear regression

analyses, investigating whether the value of local

assemblage attributes were significantly related to the

distance of sampling reaches to the stream’s mouth.

Upstream–downstream gradients in species richness,

total fish abundance and abundances of individual

species were analysed each year, to assess whether

there were inter-annual changes in spatial assemblage

patterns. Variation in the strength of such relation-

ships, as measured by the corresponding coefficients



of determination, was then related to the rainfall

variables.

Results

Environmental variation

The study encompassed a 5-year run of below-median

annual rainfall, starting in the hydrologic year of

1990/1991 and culminating in 1994/1995 with the

highest cumulative water deficit on record

()994.9 mm); annual rainfall was particularly low in

the hydrologic years of 1991/1992, 1992/1993 and

1994/1995, the last of which was the driest on record

(297.9 mm) (Fig. 1). This succession of dry years

followed an overall tendency starting in the mid

1960s, for declines in both annual and wet season

rainfall (Fig. 1). During the study, the wettest wet

seasons (rainfall > 750 mm) occurred in 1995/1996

and 1997/1998, but these were not exceptional com-

pared to others recorded in 1931–2000. Seemingly,

rainy springs (rainfall > 140 mm) recorded during the

study (1992/1993, 1995/1996 and 1996/1997) were not

exceptionally wet, though there has been a tendency

for increasing spring rainfall since about 1970.

Overall assemblage patterns

Sampling in 1991–1998 across the Torgal basin yielded

six native and two exotic freshwater fish species,

while 4–6 natives and 0–2 exotics were recorded at

individual stream reaches. There was a marked

tendency for native species richness decreasing

upstream along with distance from the stream’s

mouth (r ¼ )0.76, P < 0.05); exotics occurred only in

one upstream and two downstream reaches.

Chub and nase were the most abundant and

widespread species, occurring at all reaches and

accounting for 76% of the fish collected (Table 1).

Loach and eels also occurred at all reaches, but they

were far less abundant than the two dominant

cyprinids. Gambusia and bass were the only exotic

species, together accounting for <5% of fish catches

and occurring only at three and two reaches,

respectively. The abundance of most species showed

a tendency for declining upstream, but this was only

significant in the cases of eel and barbel (Table 1).

Changes in fish numbers

At the basin scale, there was only moderate inter-

annual variability in total fish numbers as indexed by

the CV (Table 2). Nevertheless, there was a pro-

nounced change between 1992 and 1993, when num-

bers dropped to less than half the original figures; fish

numbers recovered thereafter, declining again in

1995–96 and in 1998 (Fig. 2). The abundances of

individual species were moderately stable to highly

fluctuating, with lower CV for the most abundant

species (chub and nase) and the highest CV for the

less abundant natives (stickleback and barbel) and the

two exotics (gambusia and bass) (Table 2). Chub and

nase fluctuations followed a pattern similar to that of

total fish numbers, remaining the two most abundant

species throughout the study (Fig. 2). Loaches and

eels were always fairly abundant, though loaches

Table 1 Percentage occurrence over years

(n ¼ 8) and across reaches (n ¼ 8), and

mean annual abundances per stream

reach, of freshwater fish species sampled

in the Torgal basin (SW Portugal) in 1991–

98. Correlation (r) between fish abundance

in a reach [as log10 (X + 1)] and distance

from the reach to the mouth of the stream

is provided for each species

Species

% Occurrence Abundances Spatial trends

Years Reaches Mean ± SD (min–max) r P-value

Natives

Chub, Squalius torgalensis 100.0 100.0 65.6 ± 33.0 (26.1–128.6) 0.01 0.983

Nase, Chondrostoma almacai 100.0 100.0 50.2 ± 48.2 (7.5–162.8) )0.10 0.804

Loach, Cobitis paludica 100.0 100.0 13.1 ± 11.6 (2.6–33.9) 0.01 0.973

Eel Anguilla anguilla 100.0 100.0 10.9 ± 7.6 (1.4–22.5) )0.72 0.042

Stickleback, Gasterosteus

gymnurus

62.5 62.5 8.2 ± 9.2 (2.5–24.5) )0.20 0.638

Barbel, Barbus sclateri 87.5 62.5 0.9 ± 0.9 (0.0–2.3) )0.78 0.021

Introduced

Gambusia, Gambusia

holbrookii

50.0 37.5 2.9 ± 3.3 (0.8–6.6) )0.56 0.150

Bass, Micropterus salmoides 25.0 25.0 0.4 ± 0.4 (0.1–0.8) )0.44 0.273



declined in 1994–96 and eels in 1995–96. Stickleback

numbers varied almost 100-fold, rising steadily from

an average 0.13 to 11.1 fish 50 m)1 between 1991 and

1995, and declining thereafter to a minimum of 0.63

fish 50 m)1 in 1998. Barbel and exotic species were

always scarce, though barbel was relatively more

abundant in 1997–98 and gambusia in 1994–95.

Coefficients of variation underlined a strong ten-

dency for fish abundances to fluctuate far more at

local scale than at basin scale, although the trend for

increasing variability of individual species with

reduced average abundances remained essentially

the same (Table 2). Only eels showed a significant

trend for an upstream increase in variability (Table 2).

Rare species could not be analysed for spatial patterns

in population variability, as they were absent from

most reaches. Inter-annual changes in fish numbers at

individual stream reaches were broadly similar to that

at the basin scale, although differing in detail among

reaches (Fig. 2). At most reaches there were peak total

abundances in 1992, 1994 and 1997, while the most

pronounced decline was observed between 1992 and

1993. Eels, nase, loach and stickleback showed mod-

erate coefficients of concordance for temporal variab-

ility among reaches (W: 0.53–0.70; P < 0.001),

underlining a tendency for spatially synchronous

variations in fish abundances. Spatial concordance

was low but significant for chub, bass and gambusia

(W: 0.31–0.40; P < 0.05), but it was not significant for

barbel (W ¼ 0.23; P > 0.05).

Changes in assemblage structure

The assemblages were highly persistent at both the

basin and reach scales (T > 0.85), indicating that there

were very few changes in species composition over

time (Table 3). Fish species rank abundances were

also highly concordant, although relatively less so at

the reach than at the basin scale (Table 3). Neither

persistence nor concordance values showed any

consistent upstream–downstream pattern.

Time lag regression analysis did not revealed any

directional (r ¼ 0.11, P ¼ 0.589) or cyclic (r ¼ 0.25,

P ¼ 0.444) change in fish assemblage structure at the

basin scale. A tendency for directional change, albeit

weak, was found at a single, upstream reach (F2),

where annual assemblages appeared increasingly

different as time lags increased (Fig. 3a). A tendency

Table 2 Coefficients of variation (CV) of

annual fish catches between 1991 and 1998

in the Torgal watershed (SW Portugal), at

the basin and reach scales. Correlation (r)

between the CV in a reach and distance

from the reach to the mouth of the stream

is provided for each species recorded at all

reaches

Species Basin

Reach Up–down

Mean ± SD (min–max) n r P-value

Natives

Chub 39.3 75.3 ± 24.5 (32.3–106.0) 8 0.12 0.772

Nase 42.3 90.8 ± 23.0 (54.8–119.2) 8 0.14 0.740

Loach 55.9 110.1 ± 52.6 (72.5–227.6) 8 )0.17 0.683

Eel 33.5 66.6 ± 18.1 (50.1–102.4) 8 0.73 0.040

Stickleback 77.1 128.8 ± 30.3 (92.8–157.6) 5 – –

Barbel 138.1 181.5 ± 41.4 (138.0–233.7) 5 – –

Exotics

Gambusia 158.1 201.1 ± 49.7 (152.9–252.1) 3 – –

Bass 166.6 234.0 ± 69.1 (185.2–282.8) 2 – –

Total fish 31.6 54.1 ± 21.2 (27.6–95.6) 8 0.20 0.636

Table 3 Persistence (turnover rate) and

concordance (Kendall’s coefficient of con-

cordance) metrics for temporal changes in

fish assemblages between 1991 and 1998,

in the Torgal watershed (SW Portugal), at

the basin and reach scales. Correlation (r)

between the value of a metric in a reach

and distance from the reach to the mouth

of the stream is provided in each case

Basin

Reach Up–down

Mean ± SD (min–max) n r P-value

Persistence

Natives 0.97 0.88 ± 0.91 (0.72–1.00) 8 )0.32 0.435

Total 0.89 0.86 ± 0.91 (0.72–0.97) 8 )0.15 0.726

Concordance

Natives 0.87** 0.68 ± 0.12 (0.50* )0.84**) 8 )0.17 0.680

Total 0.90** 0.71 ± 0.11 (0.54** )0.88**) 8 )0.12 0.777

*P < 0.01; **P < 0.001.



for a cyclic pattern, with maximum differences

between annual assemblages at intermediate lags,

was found at a single downstream reach (C1; Fig. 3b).

For the remaining reaches there was no evidence from

time lag regression for either directional (P > 0.200) or

cyclic (P > 0.100) changes in assemblage structure.

Contrary to the results of the time lag regression,

PCA suggested the presence of cyclic patterns of

assemblage variation at both the basin and the reach

scales (Fig. 4). At the basin scale, the assemblage

trajectory in the first two ordination axis (57.8% of

total variation) drifted away from its initial condition

up to 1995–96, and then converged over the next

2 years to the earlier state. Changes along the first axis

were primarily related to variations in the abundances

of barbel in opposition to nase, stickleback, gambusia

and bass. The second axis was related to coordinate

changes in the abundance of common species (chub,

loach and eel), which tended to decline from 1991 to

1996, and to increase thereafter. Comparable cyclic

patterns were evident at all but the two most

upstream reaches (T4 and F2), although they were

generally less marked than at the basin scale and

differed in detail among reaches (Fig. 4). For instance,

reach C2 showed a rapid change in assemblage

structure from 1991 to 1993, followed by a slow

convergence to the earlier condition in the following

years, whereas at reach T2 the assemblage diverged

up to 1995, converged in 1996–97, and diverged again

in 1998. At the two most upstream reaches there

appeared to be a continuous drift from the initial

condition, reflecting declines in the abundances of

most species, with no signs of recovery to earlier

assemblage characteristics (Fig. 4). There were also

differences in detail among reaches on the main

species driving assemblage trajectories over time.

Nevertheless, coordinate changes in the abundance

of chub, loach, eel and, less often, nase, were related to

assemblage trajectories along the first PC axis at

several reaches.

Environmental correlates of assemblage change

Screening of environmental variables using RDA

revealed significant effects on assemblage variation

at the basin scale of variables reflecting both

sequences of dry years and wet springs. The cumu-

lative rainfall deficits (run-sums) from the long-term

median showed the strongest association with assem-

blage variation (P ¼ 0.018), whereas only nearly

significant effects were found for the corresponding

variable based on the first quartile (P ¼ 0.079). Simple

annual deficits never showed any association with

assemblage variation (P > 0.100). Spring surpluses

from the median were also significant (P < 0.030),

while surpluses from the first quartile were nearly

significant (P ¼ 0.080). Wet season surpluses were not

associated with assemblage variation (P > 0.100).

Forward selection converged on a significant model

(Monte Carlo test: P ¼ 0.0011) including both cumu-

lative deficits (P ¼ 0.018) and spring surpluses (P ¼0.044) from the median, together accounting for 49.5%

of assemblage variation over the study period. Distri-

bution of annual sample scores in the ordination space

confirmed the tendency for a cyclic pattern of assem-

blage variation, largely following the increasing rain-

fall deficit from 1991 to 1996, and its reduction

thereafter (Fig. 5). This was matched by the decline

Fig. 3 Time lag regression analysis for two stream reaches

showing directional (a) and cyclic (b) patterns of fish assemblage

variation between 1991 and 1998, in the Torgal watershed (SW

Portugal).



and subsequent recovery of chub and loach abun-

dances, while sticklebacks followed the opposite

pattern. The spring surplus defined a secondary

gradient of assemblage variation, with higher barbel

abundances associated with wetter springs, whereas

nase, eel and the two exotics favoured dryer springs.

Redundancy analysis at the scale of individual

reaches showed much weaker and inconsistent

effects of rainfall variation. At five (T1, T3, T4, C1,

C2) out of eight reaches analysed, rainfall variables

showed no significant relationships with assemblage

variation. Influences of more than one rainfall vari-

able were found at a single reach (F1), where all

annual and cumulative deficits and spring surpluses

showed significant effects (P < 0.05), although only

the annual deficit (P ¼ 0.039) from the first quartile

was retained in the forward procedure. Surplus from

the median during the wet season was the only

significant variable (P ¼ 0.038) at one reach (F2),

whereas the cumulative deficit from the median

(P ¼ 0.031) was the only significant variable at

another reach (T2). At each of these three reaches

Fig. 4 Assemblage trajectories over time (1991–98) in ordination plots derived from Principal component analysis (PCA) of fish

catches at the basin and reach scales in the Torgal watershed (SW Portugal). The percentage variance of species data and the most

correlated (>0.40) species are provided for each axis (PC).



the best RDA model accounted for 38.2–41.2% of

assemblage variation.

Effects of environmental variation on spatial patterns

Variation in rainfall conditions among years was

associated with changes in spatial patterns of assem-

blage variation. Significant (P < 0.05) trends for

decreasing species richness with increasing distance

from the stream’s mouth were recorded in 1991–93

and 1998 (r: )0.87 to )0.74; P < 0.05), but not in 1994–

97 (r: )0.65 to )0.05; P > 0.05). The strength of this

upstream–downstream gradient, as indexed by the

determination coefficient (R2), showed a strong neg-

ative relationship with the magnitude of the cumula-

tive rainfall deficit from the median (Fig. 6).

Conversely, there was no significant longitudinal

variation in overall fish catches for any individual

year (P > 0.30) but 1998, when the catches declined

upstream significantly (r ¼ )0.76, P < 0.05). Temporal

changes in this longitudinal gradient were not

significantly related to any rainfall variable.

Discussion

This study showed that Mediterranean stream fish

assemblages could be significantly affected by the

occurrence of droughts, which caused changes in

species relative abundances but had little impact on

species richness, composition (persistence) and rank

abundances (concordance). Drought effects seemed to

be cumulative across years, and were much more

evident at the catchment than at the reach scale.

Variability in assemblage structure did not show any

consistent spatial pattern, although in wet years there

was a marked tendency for a downstream increase in

species richness that disappeared in dry years.

Assemblage structure appeared resilient to the occur-

rence of droughts, as it tended to recover to its former

condition soon after the cessation of the dry period.

Besides droughts, the occurrence of rainy springs was

a significant climatic driver of assemblage change.

The main patterns revealed in this study are

unlikely to be shaped by methodological bias or

shortcomings. The study was long enough to span at

least the length of one generation for almost all the

species studied, as the assemblages were mainly

composed of small, short-lived species (Crivelli &

Britton, 1987; Cabral & Marques, 1999; Oliva-Paterna

Fig. 5 Results from a redundancy analysis relating changes be-

tween 1991 and 1998 in fish species relative abundances across

the Torgal watershed to rainfall variables. Assemblage trajectory

over time is shown by a hatched line connecting sampling years

(black circles). Dotted arrows are species and solid-line arrows

are rainfall variables. CDEF50 is the cumulative deficit from the

annual rainfall median and SSUR50 is the surplus from the

spring rainfall median. The percentage variance of species data

(%spp) and of species–environment relation (%spp–env) are

given for each axis.

Fig. 6 Relationship between the strength of the upstream–

downstream gradient in species richness and the cumulative

deficit from the median annual rainfall in the Torgal catchment

during 1991–98. The strength of the gradient was estimated each

year as the coefficient of determination of a linear regression

between the species richness in a reach and its distance to the

Mira estuary. Temporal variation in gradient strength is shown

by a hatched line connecting sampling years (black circles).



et al., 2002; Magalhaes et al., 2003) thus allowing

meaningful judgements concerning patterns of fish

assemblage variability (Grossman et al., 1990). More-

over, previous investigation has shown that the

sampling effort was sufficient to record most species

present at each sampling reach and to estimate their

relative abundances (Magalhaes et al., 2003; M.F.

Magalhaes & P. Beja, unpublished data). The 50-m

sampling reaches were probably too small to

encompass the home ranges of some species (Rodri-

guez-Ruiz & Granado-Lorencio, 1992; Prenda &

Granado-Lorencio, 1994), unlike that recommended

for studies addressing temporal variation in fish

assemblages (Grossman et al., 1990). However, errors

resulting from this potential shortcoming were prob-

ably small, because there was a low spatial turnover

of species in the rivers studied (Magalhaes et al.,

2002b), making it unlikely that species absent from a

reach could be recorded in adjacent reaches. A

potentially more serious problem was the exclusion

of age-0 fish, which was inevitable because the study

design was inadequate to sample this age group

effectively. This may have weakened the overall

association between assemblage structure and

environmental gradients, as the abundance of age-0

fish tends to be very responsive to climate conditions

(e.g. Schlosser, 1985; Schlosser et al., 2000). However,

variation in recruitment should at least be partially

reflected in the abundance of the adult population in

subsequent years (Magalhaes et al., 2003), and thus it

may be assumed that it was described indirectly in the

present study.

As recorded elsewhere (e.g. Eby et al., 2003), esti-

mates of assemblage variability increased with the

resolution of analysis, reflecting the hierarchical

nature of community dynamics (Rahel, 1990). Persist-

ence and concordance metrics were high, implying

that species composition and rank abundances stayed

essentially the same over time. Indeed, virtually every

native species was recorded each year, and there was

a core group of native species including chub, nase,

loach and eel, which remained always the most

abundant and widespread. This apparent stability in

the face of major environmental variation has already

been reported in Mediterranean-type streams (Moyle

& Vondracek, 1985) and other systems with harsh

environmental conditions (Matthews, 1986; Eby et al.,

2003; Matthews & Marsh-Matthews, 2003), probably

reflecting the role of long-term historical factors in

limiting assemblage variability (Poff & Ward, 1990;

Poff, 1997). The sequential, seasonal and highly

variable floods and drying events in Mediterranean

streams may be viewed as primary landscape filters

(sensu Poff, 1997) that reduced the species pool to

those evolutionarily adapted to cope with the pre-

vailing harsh environmental patterns (e.g. Matthews,

1987). This is supported by the observation that only

recently introduced exotics disappeared from the

basin during the study, probably because they were

less tolerant of harsh meteorological events, being

restricted to rare habitats during the dry season (large

and deep pools) and lacking life history and beha-

vioural attributes to cope with droughts and floods

(e.g. Magalhaes et al., 2002b; Bernardo et al., 2003; Eby

et al., 2003). Extinctions or major changes in rank

abundances of native species were thus unlikely

during a relatively short-term study like this, because

environmental conditions probably remained within

the range of tolerances for the species occurring in the

basin. Despite this apparent stability, however, the

abundances of both native and exotic species fluctu-

ated widely, with the highest variability for the less

abundant species, and this appeared driven to a large

extent by inter-annual variation in rainfall patterns.

The influence of rainfall on fish abundances was

probably mediated through its effects on stream flow,

primarily by determining the timing and magnitude

of extreme drought and flood events. Rainfall deficit

was one of the most influential variables, likely

reflecting variation in habitat availability and quality

during the dry season. In the hot, dry summers

characteristic of Mediterranean climates, there is a

shortage of surface waters, with fish becoming con-

fined to pool refugia and to small reaches maintaining

flowing waters, where they are under a high risk of

mortality from desiccation, predation or anoxia

(Magalhaes et al., 2002b). In particularly dry years,

the odds that fish refugia dry out before the autumn

rains increase and the size and quality of persisting

refugia decrease, thereby creating the conditions for

the occurrence of exceptionally high fish mortality

and the reduction of abundances in subsequent years

(Magalhaes et al., 2003; M.F. Magalhaes, unpublished

data). The effects of rainfall deficits appeared cumu-

lative across years, although reasons for this were

unclear. One possibility is that this might reflect a

reduced recharge of groundwaters over a sequence of

dry years, as surface waters during the dry season in



semi-arid environments may be primarily maintained

by the slow drainage of the saturated storages in the

catchment, usually local groundwater (McMahon &

Finlayson, 2003). This would imply that the amount

and quality of fish refugia in a given summer would

be conditional on the rain falling in the preceding

years. It is also possible, however, that population

mechanisms might also be at play, with for instance

measurable changes in abundance only being appar-

ent after several dry years with unusually high

mortality or reduced recruitment (Magalhaes et al.,

2003). Understanding the effects of droughts was

complicated further by the co-occurrence of other

meteorological events affecting fish assemblages. This

was the case of rainfall surpluses recorded in spring,

which likely resulted in the occurrence of floods late

in the wet season, which have the potential to cause

major mortality of both juvenile and adult fish

(Lobon-Cervia, 1996; Bernardo et al., 2003; Magalhaes

et al., 2003).

As in comparable studies elsewhere (Grossman

et al., 1998; Matthews & Marsh-Matthews, 2003),

responses to rainfall variability differed among spe-

cies, and at least three response guilds (sensu Wilson,

1999) could be identified: chub and loach declined in

abundance with increasing rainfall deficit, stickleback

increased with increasing deficit, and nase and eel

declined with increasing spring surplus. Responses by

the remaining species were difficult to assess,

although barbel increased after a succession of wet

years, thus suggesting an inverse relationship with

rainfall deficit. These contrasting responses may be at

least partly related to species-specific life history

traits, and in particular to the time of spawning

relative to drying and flooding. For instance, nase

reproduces early in the year, and this may result in

increased susceptibility to floods, but may provide the

opportunity for nase to evade stream reaches most

prone to desiccation, making them less vulnerable to

the summer drought (Magalhaes et al., 2002b, 2003).

Conversely, chub reproduces shortly before the dry-

ing-up period, when sudden floods are less likely, but

this may result in a higher risk of being trapped in

shrinking habitats during the period of receding

waters (Magalhaes et al., 2002b, 2003). The potential

processes producing the observed patterns for loach

and eel are less clear, although the former is also a late

spawner (Oliva-Paterna et al., 2002), and reductions in

eel numbers following floods were previously repor-

ted (Lobon-Cervia, 1996). Reasons for the association

of stickleback with high rainfall deficit are unknown,

but others have reported its success at exploiting

systems experiencing extreme drying (Crivelli &

Britton, 1987; Poizat & Crivelli, 1997).

Given the presence of these strong and contrasting

responses to changes in rainfall patterns, it might

seem odd the high constancy of the fish assemblages

and their quick recovery to previous condition after

the changes occurring during the sequence of dry

years. Indeed, differences in species responses to

harsh and fluctuating environmental conditions may

create temporal niche opportunities (Chesson &

Huntly, 1997), and thus compensatory changes in

the abundance of dominant species might be expected

(Collins, 2000). However, changes in dominance may

not be possible in Mediterranean streams, if too

frequent floods and droughts systematically dampen

population growth. In these circumstances, sequences

of particularly favourable years for a given species

may be too short to allow its abundance to increase

significantly, before numbers are reduced again by a

harsh event (Magalhaes et al., 2003). Therefore, if the

long-term environmental patterns remain unchanged,

time lags in population dynamics may prevent a rare

species from becoming much more common. For

instance, stickleback populations started to grow

during the sequence of dry years, but their abundance

decline sharply soon after the resuming of wetter

conditions. Seemingly, widespread and abundant

species may dominate the assemblage for extended

periods, despite moderate oscillations caused by

environmental variability. This was the case of chub

and nase, which declined in numbers after the

occurrence of droughts and spring floods, respect-

ively, but that possess the life-history attributes to

recover quickly soon thereafter (Magalhaes et al.,

2003).

Assemblage attributes measured at the scale of

individual stream reaches were more variable than

those estimated at the basin scale, but they appeared

much less influenced by changes in rainfall patterns.

Increased variability at local scales has been widely

reported, and it is usually interpreted as resulting

from spatial heterogeneity in habitats and populations

(Matthews, 1998). However, there was a tendency for

local assemblages to vary in spatial synchrony, as

observed in other Mediterranean streams (Godinho,

Ferreira & Santos, 2000), pointing to the importance of



large-scale environmental trends adding to local

habitat changes in driving assemblage structure over

time. Nonetheless, assemblages were much more

variable at some reaches than at others, indicating

that local factors may also influence their dynamics.

The degree of local assemblage variability did not

seem to follow any definite spatial pattern and, in

particular, no upstream increase in variability was

apparent, contrary to expectations developed for

temperate streams (Schlosser, 1987; Schlosser & Ebel,

1989). Lack of this longitudinal pattern may be

primarily because of the prevalence of depauperate

faunas in upstream reaches, dominated by the core

species, which tended to fluctuate little and indepen-

dently of spatial position. Conversely, the fluctuating

exotics and barbel tended to be restricted to down-

stream reaches (Magalhaes et al., 2002b), where they

added to local assemblage variability.

Despite the lack of longitudinal spatial gradients in

assemblage temporal variability, at least one upstream–

downstream trend in assemblage structure was

strongly influenced by environmental variability. In-

deed, increasing cumulative rainfall deficit seemed to

disrupt the downstream increase in species richness,

indicating that harsh drought conditions may cause the

assemblages to become less predictable relative to well-

defined spatial gradients. This confirms previous sug-

gestions that regional processes, such as droughts, can

affect longitudinal patterns in fish assemblage structure

(Horwitz, 1978; Schlosser, 1987; Paller, 1994). There-

fore, there seems to be a complex interplay between

environmental variability and the spatial and temporal

patterns of fish in Mediterranean streams, which may

be mediated by processes acting at both the catchment

and local scales.

Taken together, results from this study suggest that

the increased frequency and severity of droughts

predicted for the Mediterranean region by current

climate change models (e.g. Gibelin & Deque, 2003;

Pal et al., 2004; Santos & Miranda, 2006), may result in

significant modifications to stream fish assemblages,

with population declines or even local extinctions of

at least the species most sensitive to summer droughts

and their potential replacement by more resistant

species. Despite the observation that assemblages

were largely resistant and resilient to the occurrence

of a relatively long dry period, there was some

indication that stronger and more enduring effects

would probably have occurred if the dry spell had

lasted for a few more years, as it might be expected

under a scenario of increasing dryness. For instance,

one of the most sensitive species was the chub, which

is short-lived (<5 years) and showed recruitment

failure in very dry years (Magalhaes et al., 2003),

thereby being highly vulnerable to a dry spell lasting

for 6 years or more. The barbel appeared to be another

vulnerable species, as it is associated with large and

deep pools during the summer season, which is a rare

stream habitat in dry years (Magalhaes et al., 2002b).

In contrast, species like the stickleback may be

favoured by increasingly dry conditions (Crivelli &

Britton, 1987; Poizat & Crivelli, 1997; this study).

There is thus the potential for the ongoing changes in

drought regimes to cause a general simplification of

fish assemblages inhabiting Mediterranean streams,

which will be particularly serious for freshwater

biodiversity conservation, given the high rates of

endemism and genetic divergence of fish populations

in this region (e.g. Oberdorff et al., 1995; Cowx &

Collares-Pereira, 2000; Alves et al., 2001; Mesquita

et al., 2001, 2005). Conservation strategies to deal with

the likely increase in the frequency and intensity of

droughts thus need to be developed, so as to guaran-

tee the survival of vulnerable freshwater fish species

inhabiting Mediterranean streams.

Acknowledgments

We thank Francisco Faria, Paulo Cabrita, Joao Costa

and Conceicao Conde for their help in the field.

Critical review of the manuscript by Gary Grossman

and one anonymous referee was deeply appreciated.

Funding to this study was granted by Instituto da

Conservacao da Natureza, Agencia de Inovacao and

Fundacao para a Ciencia e Tecnologia (POCI/BIA-

BDE/56272/2004 ). I.J. Schlosser was supported by

the National Science Foundation (DEB 9973357).

References

Alves M.J., Coelho H., Collares-Pereira M.J. & Coelho

M.M. (2001) Mitochondrial DNA variation in the

highly endangered cyprinid fish Anaecypris hispanica:

importance for conservation. Heredity, 87, 463–473.

Beja P. (1996) An analysis of otter Lutra lutra predation on

the introduced American crayfish Procambarus clarkii in

an Iberian stream. Journal of Applied Ecology, 33, 1156–

1170.



Bernardo J.M., Ilheu I., Matono P. & Costa A.M. (2003)

Interannual variation of fish assemblage structure in a

Mediterranean river: implications of streamflow on the

dominance of native or exotic species. River Research

and Applications, 19, 521–532.

terBraak C.J.F. (1995) Ordination. In: Data Analysis in

Community and Landscape Ecology (Eds R.H.J. Jongman,

C.J.F. terBraak & O.F.R. van Tongeren), pp. 91–173.

Cambridge University Press, Cambridge.

terBraak C.J.F. & Smilauer P. (1998) CANOCO Reference

Manual and User’s Guide to Canoco for Windows: Software

for Canonical Community Ordination (Version 4.0). Micro-

computer Power, Ithaca, NY.

Cabral J.A. & Marques J.C. (1999) Life history, population

dynamics and production of eastern mosquitofish,

Gambusia holbrooki (Pisces, Poeciliidae), in rice fields of

the lower Mondego river Valley, Portugal. Acta

Oecologica, 2, 607–620.

Chesson P. & Huntly N. (1997) The roles of harsh and

fluctuating conditions in the dynamics of ecological

communities. American Naturalist, 150, 519–553.

Collins S.L. (2000) Disturbance frequency and commu-

nity stability in native tallgrass prairie. American

Naturalist, 155, 311–325.

Costa M.J., Costa J.L., Almeida P.R. & Assis C. (1994) Do

eel grass beds and salt marsh borders act as prefer-

ential nurseries and spawning grounds for fish? An

example of the Mira estuary in Portugal. Ecological

Engineering, 3, 187–195.

Cowx I.G. & Collares-Pereira M.J. (2000) Conservation

of endangered fish species in the face of water

resource development schemes in the Guadiana

river, Portugal: harmony of the incompatible. In:

Management and Ecology of River Fisheries (Ed. I.G.

Cowx), pp. 78–101. Fishing News Books, Blackwell

Science, Oxford.

Crivelli A.J. & Britton R.H. (1987) Life history adapta-

tions of Gasterosteus aculeatus in a Mediterranean

wetland. Environmental Biology of Fishes, 18, 109–125.

Eby L.A., Fagan W.F. & Minckley W.L. (2003) Variability

and dynamics of a desert stream community. Ecological

Applications, 13, 1566–1579.

Gasith A. & Resh V.H. (1999) Streams in Mediterranean

climate regions: abiotic influences and biotic responses

to predictable seasonal events. Annual Review of Ecology

and Systematics, 30, 51–81.

Gibelin A.L. & Deque M. (2003) Anthropogenic climate

change over the Mediterranean region simulated by a

global variable resolution model. Climate Dynamics, 20,

327–339.

Gil-Sanchez J.M. & Alba-Tercedor J. (2006) The decline of

the endangered populations of the native freshwater

crayfish (Austropotamobius pallipes) in southern Spain: it

is possible to avoid extinction? Hydrobiologia, 559, 113–

122.

Godinho F.N., Ferreira M.T. & Santos J.M. (2000)

Variation in fish community composition along an

Iberian river basin from low to high discharge: relative

contributions of environmental and temporal vari-

ables. Ecology of Freshwater Fish, 9, 22–29.

Grossman G.D., Dowd J.F. & Crawford M. (1990)

Assemblage stability in stream fishes: a review.

Environmental Management, 14, 661–671.

Grossman G.D., Ratajczak R.E. Jr, Cawford M. & Free-

man M.C. (1998) Assemblage organization in stream

fishes: effects of environmental variation and inter-

specific interactions. Ecological Monographs, 68, 395–

420.

Grossman G.D., Ratajczak R.E. Jr, Petty J.T., Hunter M.D.,

Peterson J.T. & Grenouillet G. (2006) Population

dynamics of mottled sculpin (Pisces) in a variable

environment: information theoretic approaches. Ecol-

gical Monographs, 76, 217–234.

Horwitz R.J. (1978) Temporal variability patterns and the

distributional patterns of stream fishes. Ecological

Monographs, 48, 307–321.

IPCC (2001) Climate Change 2001: The Scientific Basis.

Cambridge University Press, Cambridge.

Jones P.D., Hulme M., Briffa K.R. & Jones C.G. (1996)

Summer moisture availability over Europe in the

Hadley centre general circulation model based on the

Palmer drought severity index. International Journal of

Climatology, 16, 155–172.

Laraus J. (2004) The problems of sustainable water use in

the Mediterranean and research requirements for

agriculture. Annals of Applied Biology, 144, 259–272.

Leibold M.A., Holyoak M., Mouquet N. et al. (2004)

The metacommunity concept: a framework for

multi-scale community ecology. Ecology Letters, 7,

601–613.

Lobon-Cervia J. (1996) Response of a stream fish assem-

blage to a severe spate in northern Spain. Transactions

of the American Fisheries Society, 125, 913–919.

Magalhaes M.F., Batalha D.C. & Collares-Pereira M.J.

(2002a) Regional gradients in stream fish assemblages

across a Mediterranean landscape: relative contribu-

tions of environmental factors and spatial structure.

Freshwater Biology, 47, 1015–1031.

Magalhaes M.F., Schlosser I.J. & Collares-Pereira M.J.

(2003) The role of life history in the relationship

between population dynamics and environmental

variability in two Mediterranean stream fishes. Journal

of Fish Biology, 63, 300–317.

Magalhaes M.F., Beja P., Canas C. & Collares-Pereira M.J.

(2002b) Functional heterogeneity of dry-season fish

refugia across a Mediterranean catchment: the role of



habitat and predation. Freshwater Biology, 47, 1919–

1934.

Matthews W.J. (1986) Fish faunal structure in an Ozark

stream: stability, persistence and a catastrophic flood.

Copeia, 1986, 388–397.

Matthews W.J. (1987) Physicochemical tolerance and

selectivity of stream fishes as related to their geo-

graphic ranges and local distributions. In: Community

and Evolutionary Ecology of North American Stream Fishes

(Eds W.J. Matthews & D.C. Heins), pp. 111–127.

University of Oklahoma Press, Norman, OK.

Matthews W.J. (1998) Patterns in Freshwater Fish Ecology.

Chapman & Hall, New York, NY.

Matthews W.J. & Marsh-Matthews E. (2003) Effects of

droughts on fish across axes of space, time and

ecological complexity. Freshwater Biology, 48, 1232–

1253.

McMahon T.A. & Finlayson B.L. (2003) Droughts and

anti-droughts: the low flow hydrology of Australian

rivers. Freshwater Biology, 48, 1147–1160.

Mesquita N., Hanfling B., Carvalho G.R. & Coelho M.M.

(2005) Phylogeography of the cyprinid Squalius araden-

sis and implications for conservation of the endemic

freshwater fauna of southern Portugal. Molecular

Ecology, 14, 1939–1945.

Mesquita N., Carvalho G., Shaw P., Crespo E. & Coelho

M.M. (2001) River basin-related genetic structuring in

an endangered fish species, Chondrostoma lusitanicum,

based on mtDNA sequencing and RFLP analysis.

Heredity, 86, 253–264.

Millan M.M., Estrela M.J. & Miro J. (2005) Rainfall

components: variability and spatial distribution in a

Mediterranean area (Valencia Region). Journal of

Climate, 18, 2682–2705.

Moye L.A., Kapadia A.S., Cech I.M. & Hardy R.J. (1988)

Theory of runs with applications to drought predic-

tion. Journal of Hydrology, 103, 127–137.

Moyle P.B. & Vondracek B. (1985) Persistence and

structure of the fish assemblage in a small California

stream. Ecology, 66, 1–13.

Oberdorff T., Guegan J.F. & Hugueny B. (1995) Global

scale patterns of fish species richness in streams.

Ecography, 18, 345–352.

Oliva-Paterna F.J., Torralva M.M. & Fernandez-Delgado

C. (2002) Age, growth and reproduction of Cobitis

paludica in a seasonal stream. Journal of Fish Biology, 60,

389–404.

Pal J.S., Giorgi F. & Bi X. (2004) Consistency of recent

European summer precipitation trends and

extremes with future regional climate projections.

Geophysical Research Letters, 31, L13202 doi:10.1029/

2004GL019836.

Paller M.H. (1994) Relationships between fish assem-

blage structure and stream order in South Carolina

Coastal Plain streams. Transactions of the American

Fisheries Society, 123, 150–161.

Paulo A.A., Pereira L.S. & Matias P.G. (2003) Analysis

of local and regional droughts in southern Portugal

using the theory of runs and the standardized

precipitation index. In: Tools for Drought Mitigation

in Mediterranean Regions (Eds G. Rossi, A. Cancelliere

& L.S. Pereira), pp. 55–78. Kluwer Academic Pub-

lishers, Dordrecht.

Poff N.L. (1997) Landscape filters and species traits:

towards mechanistic understanding and prediction in

stream ecology. Journal of the North American Bentho-

logical Society, 16, 391–409.

Poff N.L. & Ward J.V. (1990) Physical habitat template of

lotic systems: recovery in the context of historical

pattern of spatiotemporal heterogeneity. Environmental

Management, 14, 629–645.

Poizat G. & Crivelli A.J. (1997) Use of seasonally flooded

marshes by fish in a Mediterranean wetland: timing

and demographic consequences. Journal of Fish Biology,

51, 106–119.

PrendaJ.&Granado-LorencioC.(1994)Estimasdelespacio

vital y calidad del habitat a lo largo del Invierno en tres

especies de peces (Cyprinidae) de un rio de regimen

Mediterraneo. Donana, Acta Vertebrata, 21, 61–77.

Rahel F.J. (1990) The hierarchical nature of community

persistence: a problem of scale. American Naturalist,

136, 328–344.

Rodrigues R., Brandao C. & Alvares T. (1998) Qual o Grau

de Excepcionalidade das Cheias Ocorridas no Inıcio do ano

Hidrologico de 1997/98? Instituto da Agua, Lisboa,

Portugal.

Rodriguez-Ruiz A. & Granado-Lorencio C. (1992)

Spawning period and migration of three species of

cyprinids in a stream with Mediterranean regimen

(SW Spain). Journal of Fish Biology, 41, 545–556.

Santos F.D. & Miranda P. (eds). (2006) Alteracoes

Climaticas em Portugal: Cenarios, Impactos e Medidas de

Adaptacao. Gradiva, Lisboa, Portugal.

Schlosser I.J. (1985) Flow regime, juvenile abundance,

and the assemblage structure of stream fishes. Ecology,

66, 1484–1490.

Schlosser I.J. (1987) A conceptual framework for fish

communities in small warmwater streams. In: Commu-

nity and Evolutionary Ecology of North American Stream

Fishes (Eds W.J. Matthews & D.C. Heins), pp. 17–24.

University of Oklahoma Press, Norman, OK.

Schlosser I.J. & Ebel K.K. (1989) Effects of flow regime

and cyprinid predation on a headwater stream.

Ecological Monographs, 59, 41–57.



Schlosser I.J., Johnson J.D., Knotek W.L. and Lapinska M.

(2000) Climate variability and size-structured interac-

tions among juvenile fish along a lake-stream gradient.

Ecology, 81, 1046–1057.

Schonwiese C.D. & Rapp J. (1997) Climate Trend Atlas of

Europe Based on Observations 1891–1990. Kluwer Aca-

demic Publishers, Dordrecht, The Netherlands.

Siegel S. & Castellan N.J. (1988) Nonparametric Statistics

for the Behavioural Sciences. McGraw-Hill International,

New York, NY.

Smith K.G. & Darwall W.R.T. (2006) The Status and

Distribution of Freshwater Fish Endemic to the Mediterra-

nean Basin. IUCN – The World Conservation Union,

Gland, Switzerland.

Wilson J.B. (1999) Guilds, functional types and ecological

groups. Oikos, 86, 507–522.

Yevjevich V. (1967) An objective approach to definitions

and investigations of continental hydrologic droughts.

Colorado State University Hydrology Paper, 23, 1–18.

(Manuscript accepted 26 March 2007)



Effects of multi-year droughts on fish assemblages of seasonally drying Mediterranean streams

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