PRIMARY RESEARCH PAPER
Assessing the effects of sewage effluents in a Mediterraneancreek: fish population features and biotic indices
Blanca Figuerola • Alberto Maceda-Veiga •
Adolfo De Sostoa
Received: 26 December 2011 / Revised: 13 April 2012 / Accepted: 24 April 2012
� Springer Science+Business Media B.V. 2012
Abstract Sewage effluents are one of the main
anthropogenic stressors in Mediterranean rivers. The
establishment of a cause–effect relationship is hin-
dered in natural systems by the existence of con-
founding factors (i.e. biotic interactions). Here we
analysed the effects that anthropogenic stressors have
on a mono-specific fish community (Iberian redfin
barbel population, Barbus haasi) inhabiting the north-
ern edge of its distribution range. In Spring 2004, a
total of 40 consecutive sampling sites were surveyed
in Vallvidrera creek, and 1,331 specimens were
measured and weighed. A principal component anal-
ysis was performed to synthesize the information
provided by 22 environmental variables. Analysis of
variance, bivariate correlation analyses and multiple
linear regressions were then used to determine the
influence of the environmental gradients built (water
quality, hydromorphology, woods and macrophytes,
and degree of silting) on fish population features (fish
size, body condition status, density and biomass).
The findings revealed that water quality was the
most significant environmental gradient for this fish
population. In particular, fish density decreases and
fish length increases in those sites exposed to sewage.
Additionally, our results showed the best body con-
dition of those specimens inhabiting fast flow reaches
which confirms the rheophilous condition of B. haasi.
However, these findings were unnoticed for the
current version of the index of biotic integrity using
fish as bioindicators in Catalonia. Resource managers
need to refine diagnostic tools in order to detect subtle
deleterious changes on fish communities before they
become evident at population scale. Conservation
measures should be focused in these small streams in
where the best preserved native fish populations
usually inhabit. This study suggests the need to change
water management policies in Mediterranean rivers to
improve the water quality of sewage effluents and
increase the dilution power of these rivers.
Keywords Mediterranean creek � Sewage effluents �Effects � Iberian redfin barbel
Introduction
Mediterranean rivers have suffered a long history of
anthropogenic impact events, which have conditioned
the conservation status of their native fish fauna. Fish
species are also shaped by the harsh climatic con-
straints of the Mediterranean area in which cycles of
droughts and floods alternate (Gasith & Resh, 1999;
Magalhaes et al., 2007; Verdiell et al., 2007; Beche
Handling editor: M. Power
B. Figuerola (&) � A. Maceda-Veiga � A. De Sostoa
Department of Animal Biology & Biodiversity Research
Institute (IRBIO), Faculty of Biology, University of
Barcelona, 08028 Barcelona, Spain
e-mail: [email protected]
123
Hydrobiologia
DOI 10.1007/s10750-012-1132-y
et al., 2009). Both anthropogenic and natural pres-
sures, which may also exacerbate the consequences of
the former, have converted Mediterranean rivers into
one of the most endangered ecosystems in the world
(Hooke, 2006; Magalhaes et al., 2008).
Native fish species from Mediterranean streams
have adapted their life-histories (e.g. fast growth, early
maturity and multiple spawning) to face the high-
hydrological variability of these systems (e.g.
Fernandez-Delgado & Herrera, 1995; Vinyoles et al.,
2010). In this regard, the most restrictive conditions
appear in small intermittent streams, because fish need
to develop the ability to resist floods and survive in
isolated pools during drought (Aparicio & de Sostoa,
1998; Pires et al., 1999; Soriguer et al., 2000;
Magalhaes et al., 2002). Pools act as a population
reservoir, allowing fish recolonization when flow is re-
established (Sedell et al., 1990; Magalhaes et al., 2007;
Pires et al., 2010). Nevertheless, flow variations are
still the main natural conditioners of the abiotic (i.e.
channel morphology) and biotic (i.e. fish assemblages)
structure of Mediterranean rivers (Magalhaes et al.,
2007; Boix et al., 2010). Natural variations in fish
assemblages and in the structure of fish populations
have been reported in these systems within seasons but
also inter-annually (Magalhaes et al., 2007; Beche
et al., 2009).
With regard to anthropogenic alterations, the input
of sewage treatment plants plays a crucial role in the
water quality and flow of Mediterranean rivers, in
particular in those streams with the most restrictive
hydrological conditions during droughts (Prat &
Munne, 2000; Maceda-Veiga et al., 2009; Damasio
et al., 2011). Natural flow reduction added to water
abstractions may diminish the dilution capacity of
streams, and the effects of pollutants on fish fauna
may be exacerbated (Gasith & Resh, 1999; Prat &
Munne, 2000; Pires et al., 2010). This is particularly
important in areas with a high degree of endemicity,
such as the Iberian Peninsula, or when it affects fish
populations with great conservation value (Sostoa,
1990; Doadrio, 2001; Maceda-Veiga et al., 2010a).
NE Spain has a long history of anthropogenic
alterations, such as the modification of the morphol-
ogy of the channels, abstractions, changes in land
use and industrial and municipal sewage discharges
(Prat & Ward, 1994). In addition, its native fish
fauna have declined greatly in recent decades
(Maceda-Veiga et al., 2010a).
Barbus haasi (Mertens, 1925) is a medium-sized
cyprinid that inhabits the headwaters and middle
reaches of Mediterranean Spanish Rivers from the
Llobregat to the Turia basins. It is listed as rare (R) in
the Redbook of Spanish Vertebrates (Doadrio, 2001).
It also appears as vulnerable in international and
national IUCN criteria (Doadrio, 2001; Crivelli,
2005). Unlike main rivers, in which introduced
species tend to dominate fish assemblages, some
tributaries still preserve native fish fauna populations
in good conservation status. Previously published
studies have addressed the life-history, home-range
and habitat preferences of this population of B. haasi
(Grossman & Freeman, 1987; Grossman & de
Sostoa, 1994; Aparicio & de Sostoa, 1999). Never-
theless, the relationships between fish abundance, fish
condition and anthropogenic stressor gradients
related to habitat and water quality were not assessed.
This fish community consists solely of the Iberian
redfin barbel (B. haasi), which is an advantage for
establishing relationships between fish species and
environmental factors, since biotic interactions are
minimized (Gasith & Resh, 1999). In addition, this
population is at the edge of the northern range of its
distribution (Vallvidrera creek, Llobregat basin),
which may condition its susceptibility to anthropo-
genic impacts (Gasith & Resh, 1999; Maceda-Veiga
& de Sostoa, 2011).
Understanding the response of native fish fauna
to anthropogenic stressors is crucial to guarantee-
ing their conservation and to refining the use of
fish as bioindicators (Karr, 1991; Schiemer, 2000;
Vila-Gispert et al., 2002; Oliva-Paterna et al.,
2003; Benejam et al., 2008; Maceda-Veiga et al.,
2010b; Maceda-Veiga & de Sostoa, 2011). Fur-
thermore, environmental risk assessment studies
using fish as bioindicators have been poorly
interlinked with the results obtained by other
biotic indices (e.g. riparian coverage), which may
increase their value as diagnostic tools (Benejam
et al., 2008). Here, we studied the population of
B. haasi in a small Mediterranean creek (Vallvid-
rera), in order to determine the effects that habitat
and water quality have at the population (size–age
structure, fish density and biomass) and individual-
organism level (body condition). In addition, we
compared the suitability of two biotic indices
(IBICAT and QBR) as diagnostic tools in small
Mediterranean streams.
Hydrobiologia
123
Materials and methods
Study area
This reach surveyed is located in Vallvidrera Creek
(418260N, 2�030E) in Llobregat basin within the Col-
lserola Natural Park, close to Barcelona city in
Catalonia (northeast of Spain, Fig. 1). Vallvidrera
Creek is a first-order tributary of the River Llobregat.
It is 4 km in length and has a mean stream width of
1.60 m. The highest flows usually occur in spring and
the lowest in summer, when some stretches of the
stream can lose connectivity and fish remain in isolated
pools. Other than rainfall, water flow is influenced by
the effluent from an activated sludge sewage treatment
plant located close to the source of Vallvidrera Creek
which represents over the 50–100 % of rivers’ flow
(rainy and drought conditions, respectively), high-
substrate permeability and water extractions for agri-
cultural purposes (Murria, 2003; Maceda-Veiga et al.,
2009). In addition, the stream regularly receives effluent
inputs from the sewage collectors of nearby urban
areas (physicochemical parameters are as follows:
ammonium = 21.5 mg/l, phosphates = 11.2 mg/l and
conductivity = 1,541 lS/cm, ACA 2006). The water
quality of Vallvidrera creek varies as follows:
Fig. 1 Reach surveyed
each 50 m in the Vallvidrera
creek (Llobregat basin)
(Catalonia, NE Spain)
Hydrobiologia
123
temperature ranges between 4 �C in winter and 20 �C
in summer, dissolved oxygen is always close to
saturation, with the exception of isolated pools during
the summer, when minimum values of 3–6 mg/l were
recorded and conductivity ranges between 800 and
1,100 lS/cm (Maceda-Veiga et al., 2009). The current
ichthyofauna of this stream comprises only B. haasi,
but the European eel (Anguilla anguilla) was also
found some years ago (Aparicio & de Sostoa, 1999).
Surveys were performed covering the whole 2 km
section of this creek inhabited by fish. The area
surveyed was divided in 50 m long consecutive
reaches using blocking nets.
Environmental data
Prior to fish sampling, physicochemical water quality
parameters were measured at each sampling site. A
digital multiparametric YSI� 556 MPS probe was
used for conductivity (lS/cm), dissolved oxygen (O2,
mg/l), temperature (�C) and pH. The colorimetric test
kit VISOCOLOR� was used to determine ammonia
(NH3, mg/l), nitrite (NO2, mg/l), nitrate (NO3, mg/l)
and phosphate (PO4, mg/l) levels. After sampling fish,
the hydro-morphological variables of the river (cur-
rent velocity, width and depth) were measured along
transects set perpendicular to the flow at 20 m
intervals. We then calculated the total area surveyed
and the mean depth. To assess the relative proportion
of different substrate types (%), we visually catego-
rized the substrate type in: bedrock (exposure of the
consolidated rock), rocks and pebbles (64–256 mm),
gravel and sand (0.06–64 mm), silt (0.06–0.004 mm)
and detritus (organic material including the bodies or
fragments of dead organisms). We also characterized
the potential refuges for fish and other habitat diversity
features: structural refugia (%), caves (%), aerial
coverage (%), aquatic macrophytes (%), submerged
riparian vegetation (%) and woods (%). To further
characterize habitat, we also applied the riparian
vegetation quality index (QBR) (Munne et al., 1998).
The QBR is a summation index of four parts: total
riparian cover, cover structure, cover quality and
naturalness of river channel. Each part is calculated
independently, and the individual score of each part
cannot be either negative or higher than 25. Then, the
final score is the sum of these four parts and ranked
between 0 and 100, in where 100 indicate the highest
degree of conservation.
Fish sampling
Fish were captured with a portable electro-fishing unit,
which generates a voltage up to 400 V, 2 A, and uses
the single-pass electric fishing method (CEN, 2003).
Block nets were used to avoid fish movements
between reaches. All fish were identified at species
level and counted, and a subsample of 50 randomly
selected fish (when it was possible) was measured. To
avoid management stress, fish were anaesthetized with
MS-222 (3-aminobenzoic acid ethyl ester; Sigma-
Aldrich�) and were returned alive to the river after
examination. We determined the fork length (FL, mm)
and total wet weight (mg). All specimens were
examined in situ by the naked eye for external
macroscopic parasites or anomalies. Relative fish
density was calculated, using the total number of fish
captured, the area surveyed and the fishing time
(CPUE, Capture per Unit of Effort). On the basis of the
fish captured at each sampling site, its ecological
status was determined, following the index of biotic
integrity developed for Catalan Rivers (IBICAT)
(Sostoa et al., 2003). The IBICAT’s methodology, as
other IBIs, consists of comparing the ecological status
of two body masses (the reference and the tested one)
based on selected features named metrics (e.g. fish
density, % introduced species, % of insectivorous
species, % of tolerant species, etc.) (Karr, 1991;
Sostoa et al., 2003; Pont et al., 2007).
Data analyses
Normality and homoscedasticity were tested by the
Shapiro–Wilk and Levene tests, respectively. Data
transformation did not allow normality and homosce-
dasticity to be attained. To avoid redundancy in further
analyses, the bivariate relationship between environ-
mental variables was checked by Spearman’s corre-
lation coefficient (one variable for each pair-wise with
r [ 0.7 was removed). Principal component analysis
(PCA) was then applied as an indirect ordination
technique to describe the main sources of variation and
relationship among the selected environmental vari-
ables. PCA reduces the dimensionality of the envi-
ronmental variables to a few principal synthetic
gradients. The ‘‘varimax’’ rotation method was used
to increase the interpretation of axes; and the number
of PCA axes examined was determined by Kaiser’s
Hydrobiologia
123
rule, which states that the minimum eigenvalue should
be 1 when correlation matrices are used (Legendre &
Legendre, 1998). The PCA scores for each gradient
were used for regression analysis and for grouping
sampling sites based on their factor loadings. The
Kruskal–Wallis test was used to compare between fish
features (FL, weight, condition factor and density) and
groups. The somatic condition of B. haasi was
calculated with Fulton’s condition factor (e.g. Clavero
et al., 2009): K = W � 105 � FL3, where W is the wet
weight (in g) and FL is the FL (in mm). The U Mann–
Whitney test was used for paired comparisons. The
bivariate relationship between environmental gradi-
ents (i.e. PCA scores) and fish features was checked by
Spearman’s correlation coefficient. To further
describe the relationship between biological and
environmental data, multiple linear regression models
were performed to test the relationship between fish
density, biomass and the average length, weight
(W) and condition factor (CF) of fish and the
synthetic PCA gradients. We followed a manually
conducted backward stepwise procedure to construct
minimum adequate models (MAM). Retention of
independent variables in the models relied on
statistical significance (P \ 0.05), a more restrictive
criterion than information criteria (e.g. Akaike
information criterion), which tend to leave more
independent variables in final models (Crawley,
2002; Clavero et al., 2009).
All analyses were performed using the R package
(R Development Core Team, 2010) and the library
statistics, lawstat (Nogouchi et al., 2009) and psych
(Revelle, 2010), assuming the alternative hypothesis
at P \ 0.05.
Results
A total of 40 sampling sites were surveyed, and
fish were captured at all of them. 1,331 specimens
were measured and weighed, and 22 environmental
variables were determined (Table 1). The strongest
correlations were detected between the following envi-
ronmental variables: nitrates with nitrites (r = 0.712,
P \ 0.001), pH with alkalinity (r = 0.705, P \0.001),
gravel and sand with rocks and pebbles (r = 0.734,
P \ 0.001), structural refuge with coves (r = 0.707,
P \ 0.001) and submerged riparian coverage with
natural woods (r = 0.701, P = 0.01).
Environmental gradients
The PCA produced four significant axes, which
explained 70.13 % of variation in our environmental
data set. The suitability of this analysis was deter-
mined by the Kaiser–Meyer–Olkin test (KMO =
0.703) (Table 2). PC1 accounted for 25.52 % of
Table 1 Mean, standard deviation and minimum maximum
values of the environmental and biological variables recorded
at the sampling sites surveyed, without removal of redundant
environmental variables
Mean ± SD Range
(min–max)
Biological data
Density (ind/ha) 6781.9 ± 4133 178.2–25163.4
Biomass (kg/ha) 202308 ± 167401 5542–698440
FL (mm) 108.30 ± 29.01 23–263
W (g) 32.49 ± 24.13 0.2–314
CF 0.138 ± 0.02 0.12–0.13
Biotic indices
QBR 33.62 ± 11.20 10–55
IBICAT 5 ± 0 5–5
Environmental data
Ammonia (mg/l) 0.355 ± 0.255 0–1.50
Nitrite (mg/l) 0.162 ± 0.167 0–0.50
Nitrate (mg/l) 7.79 ± 5.49 5–18
Phosphate (mg/l) 0.62 ± 0.23 0.20–1.20
Alkalinity (� KH) 12.90 ± 0.42 10–15
pH 7.55 ± 0.23 7.20–8.00
Depth (m) 0.941 ± 0.16 0.55–1.35
Current velocity (m/s) 0.272 ± 0.18 0.07–1.00
Temperature (�C) 12.04 ± 2.86 8.20–17.00
Conductivity (lS/cm) 936.9 ± 229.99 534–1148
Bedrock (%) 1.55 ± 2.86 0–10
Rocks and pebbles (%) 14.5 ± 8.21 0–25
Gravel and sand (%) 73.30 ± 8.50 61–93
Silt (%) 2.85 ± 1.25 1–6
Detritus (%) 1.00 ± 0.32 0–2
Structural refugia (%) 7.28 ± 3.84 2–20
Caves (%) 6.55 ± 5.54 0–15
Aerial coverage (%) 51 ± 42 10–90
Aquatic macrophytes
(%)
3.60 ± 2.34 1–6
Submerged riparian
vegetation (%)
15.88 ± 12.32 10–35
Woods (%) 18.00 ± 15.54 10–30
Hydrobiologia
123
variation and loaded mainly on physico-chemical
water parameters and the quality of riparian coverage.
PC2 explained 20.71 % of variation and included
hydromorphological variables (current velocity, struc-
tural refugia, % bedrock). PC3 accounted for 13.51 %
of variation and loaded on the percentage of woods
and macrophytes; whereas PC4, accounting for
10.39 % of variation, included the degree of silting.
Sampling sites were grouped, based on their scores
in the two main environmental stressor gradients
(Fig. 2). The effects of the effluents from the sewage
treatment plant and nearby urban collectors condi-
tioned the location of Groups I and III in the PC1
gradient. Group II mainly reflected reaches from
headwaters with high-current velocity and low-detri-
tus deposition.
Fish population features and biotic indices
The characteristics of specimens sampled at each
group (I, II, III) were provided in Table 3. All
specimens differed in the average FL, W) and body
CF between groups. In particular, individuals from
Group III had, significantly, the lowest values in FL
and W (P \ 0.001). There were significant differences
between CF and all groups (P \ 0.001). Fish density
decreased significantly in locations from Group III
(P \ 0.001), but no significant differences were
detected in biomass (P = 0.448).
Groups also differed in the quality of the riparian
coverage (QBR). Locations in Group II have signif-
icantly the highest values in this index (P \ 0.001). In
contrast, as all sampling sites received the same score,
based on IBICAT, there were no significant differ-
ences between groups or a positive correlation
(P [ 0.05).
Environmental gradients and fish population
features
The multiple linear regression analyses explained 62.9
and 38.9 % of the variation in Iberian redfin barbel
density and the specimen’s length. Both models are
selected, as a significant independent variable, PC1
(b = -0.782; P \ 0.001; b = 0.618; P \ 0.001,
respectively). Bivariate correlation analyses also
revealed the high-significant relationship between
PC1 and fish density, and fish size compared to the
others (Table 4). After removal of the non-significant
factors (MAM), the relationship between these bio-
logical data and PC1 revealed that Iberian redfin
barbels tended to be scarce and specimens big in
Table 2 Main environmental gradients (PC1, PC2, PC3 and
PC4) defined in the study area by means of PCA
Environmental variable PC1 PC2 PC3 PC4
Ammonia 0.63 0.09 0.23 -0.22
Nitrite 0.51 0.28 0.54 0.14
pH 0.82 -0.05 -0.18 0.33
Average depth 0.64 0.14 0.33 0.09
Average water current -0.01 0.74 -0.07 0.49
Water temperature -0.82 -0.18 0.18 -0.42
Conductivity 0.90 0.04 -0.05 0.12
Bedrock 0.12 0.82 0.09 -0.11
Gravel and sand -0.46 -0.51 -0.36 -0.13
Silt 0.24 0.15 0.13 0.77
Detritus -0.02 -0.82 0.28 -0.07
Structural refuge 0.03 0.72 0.40 0.24
Macrophytes -0.19 0.15 0.72 0.06
Submerged riparian
coverage
-0.08 0.61 0.46 0.45
Natural Woods 0.25 -0.24 0.72 -0.05
QBR 0.68 -0.24 0.05 -0.40
The greatest factor loadings in absolute value are marked in
bold
Fig. 2 Sampling scores on the first two component axes: PC1
mainly loaded on water quality and riparian coverage (25.52 %)
and PC2 mainly loaded on hydromorphology (20.71 %).
Groups (I, II and III) were based on the proximity of sampling
sites in the PCA
Hydrobiologia
123
locations with high conductivity, ammonia and ripar-
ian quality index (QBR) values, and low-water
temperature, depth and pH (R2 = 0.60; b = -0.782;
P \ 0.001; R2 = 0.36; b = 0.618; P \ 0.001; respec-
tively) (Fig. 3). A more detailed analysis of fish
lengths showed that the minimum FL of specimens
increased along PC1 (R2 = 0.37; b = 0.622;
P \ 0.001) (Fig. 3). In contrast, no significant rela-
tionship was found between the maximum FL and
habitat gradients (R2 = 0.10; P = 0.100). The same
habitat gradient (PC1) was selected as significant in
the case of fish weight (R2 = 0.32; b = 0.511;
P \ 0.001). In addition, the model explained 40.9 %
of the variation in the body CF, and the significant
variable selected was PC2 (b = 0.598; P \ 0.001).
The regression between the body CF and PC2
excluding the rest showed that fish had a good
condition in locations with high-water current, struc-
tural refuges, submerged riparian coverage and low
detritus (R2 = 0.34; b = 0.598; P \ 0.001).
Discussion
Environmental parameters
Physico-chemical water parameters indicated a
decrease in water quality downstream. The poor water
quality was mainly due to the increase in ionic levels
and organic pollution, denoted by an increase in
conductivity, ammonia, nitrite, nitrate and phosphate
values downstream of the sewage inputs. The presence
of reduced forms of nitrogenous compounds (ammo-
nia, nitrites) shows the input of organic sewage
effluents without treatment or from an inefficient
sewage treatment plant (e.g. Welch & Lindell, 1992;
Mariz, 2011). The eutrophication detected may also be
responsible for the proliferation of algae (i.e.Ta
ble
3M
ean
,st
and
ard
dev
iati
on
,m
axim
um
and
min
imu
mv
alu
esfo
rb
iolo
gic
alp
aram
eter
san
db
ioti
cin
dic
esp
rov
ided
by
gro
up
sb
ased
on
PC
Asc
ore
s
Gro
up
IG
rou
pII
Gro
up
III
Sta
tist
ics
Pv
alu
e
Mea
n±
SD
(min
–m
ax)
Mea
n±
SD
(min
–m
ax)
Mea
n±
SD
(min
–m
ax)
KW
Den
sity
(in
d/h
a)4
32
0±
13
56
(33
61
–5
27
9)a
41
89
±3
84
6(1
78
–1
56
86
)a1
45
34
±5
41
9(6
28
6–
25
16
3)b
18
.17
\0
.01
Bio
mas
s(k
g/h
a)1
88
98
3±
54
47
6(1
50
46
2–
22
75
03
)a1
79
37
7±
15
10
90
(55
42
–5
03
97
2)a
26
91
78
±2
14
41
9(3
07
32
–6
98
44
0)a
1.6
10
.45
FL
(mm
)1
20
.51
±4
.32
(36
–2
21
)a1
25
.4±
22
.72
(23
–2
62
)a8
4.7
6±
23
.72
(23
–2
63
)b1
94
.19
\0
.01
W(g
)4
3.7
8±
1.1
7(0
.7–
19
6.5
)a4
2.6
6±
24
.37
(0.2
–3
14
)a1
8.2
4±
10
.04
(0.2
–2
42
)b2
17
.47
\0
.01
CF
0.2
0±
0.0
8(0
.11
–1
.95
)a0
.15
±0
.01
(0.0
1–
1.9
3)b
0.1
3±
0.0
1(0
.05
–0
.6)c
31
0.9
9\
0.0
1
QB
R1
0±
0(1
0–
10
)a3
8.3
9±
9(1
5–
55
)b2
5±
0(2
5–
25
)a1
8.4
0\
0.0
1
IBIC
AT
5±
0(5
–5
)a5
±0
(5–
5)a
5±
0(5
–5
)a0
.00
1.0
0
Th
ele
tter
s(a
,b
,c)
con
sid
erth
atg
rou
ps
wer
eh
om
og
eneo
us
atP
\0
.05
Table 4 Correlation matrix (Spearman’s coefficient) of the
fish data and the environmental gradients built in PCA
PC1 PC2 PC3 PC4
Fish density -0.793** -0.227 -0.003 -0.016
Fish biomass -0.399* -0.207 0.130 0.068
FL 0.555* 0.112 0.141 0.016
W 0.464* 0.152 0.223 0.074
CF 0.461* 0.198 0.051 0.097
* P \ 0.05, ** P \ 0.001
Hydrobiologia
123
Cladophora). This fact may also explain the increase
in pH in those reaches due to the photosynthetic
activity of these algae mass.
Previous biomonitoring studies of macroinverte-
brates and algae also revealed a medium level of
organic pollution in this creek (Cambra et al., 2000;
Murria, 2003). In addition, these studies did not show
significant variations in water chemical parameters
between years, which indicate that aquatic fauna are
chronically exposed to this pollution gradient. In this
regard, chronic pollution events have been reported to
enhance the susceptibility of fish to disease (Maceda-
Veiga et al., 2009).
Fish population and environmental parameters
The effects of chronic pollution on native fish
populations have been poorly studied in small Med-
iterranean rivers (Vila-Gispert et al., 2002; Oliva-
Paterna et al., 2003; Damasio et al., 2007; Maceda-
Veiga et al., 2010b). Our findings revealed changes in
population structure, fish density and body condition
along the pollution gradient analysed. A significant
decrease in fish density was found in sites with high-
anthropogenic pressures (e.g. pollution). Neverthe-
less, fish density fluctuations can be common in
Mediterranean fish populations due to natural climatic
Fig. 3 Relationship between redtail barbel features (density, size and body condition) and the habitat gradient selected in the minimum
adequate model analyses (MAM)
Hydrobiologia
123
constraints (i.e. floods–drought) and it is often difficult
to discriminate between the factors affecting fish
populations (Magalhaes et al., 2007; Beche et al.,
2009).
Biases in population structure have been reported in
those fish populations subjected to strong environ-
mental stress (Adams et al., 1992; Gafny et al., 2000).
The life-history traits of native fish have evolved to
cope with natural climatic constraints, and these traits
may also be an advantage for surviving a pollution
event. Fish population is skewed to small sizes in
polluted sites, but native fish species from such
streams tend to have high fertility and multiple-
spawning events and reach sexual maturity at a
younger age (Herrera, 1991; Fernandez-Delgado &
Herrera, 1995; Vinyoles et al., 2010). Sexual maturity
is reached at 45 and 98 mm FL in males and females,
respectively, in this stream and fish lengths recorded
were always over these values in all sampling sites
(Aparicio & de Sostoa, 1998). Nevertheless, it is worth
nothing that small females have a less fecundity than
the larger ones (Aparicio & de Sostoa, 1998).
Sewage effluents usually contain a mixture of
xenobiotics (e.g. drugs, heavy metals, polychlorinated
biphenyls, etc.) that are responsible for a wide range of
physiological dysfunctions on fish fauna (e.g. geno-
toxic effects, tissue damage, metabolic anomalies,
endocrine disruption) (Di Giulio & Hinton, 2008).
Indeed, endocrine disruptors may shift the population
sex-ratio towards the females which are larger than
males, and this fact could be an alternative explanation
to this tendency detect in our study (Devlin &
Nagahama, 2002). However, the sexual dimorphism
of this species is not evident out of breeding season
(Aparicio & de Sostoa, 1998). Above all, the persis-
tence of this population may be guaranteed whether
pollutant levels are maintained within the tolerance
range of this fish species. B. haasi is considered
intolerant to poor water quality, and these conditions
may increase its susceptibility to disease, as has already
been reported (Maceda-Veiga & de Sostoa, 2011). In
any case, it should be pointed out that hydromorpho-
logical parameters (i.e. water level) seem to be a
greater fish size conditioner than chemical parameters
alone, particularly in small streams (Aparicio & de
Sostoa, 1998; Gasith & Resh, 1999; Pires et al., 2010).
In this regard, large fish tend to be more susceptible to
diseases in these streams in low-water flow conditions
(i.e. drought) (Maceda-Veiga et al., 2009).
General condition indices provide a first approach to
determining the health status of a fish population (Van
der Oost et al., 2003). Nevertheless, as reported in
similar studies (Van der Oost et al., 1998), the CF values
of fish exposed to sewage effluents in our study were not
significantly different from those registered in the
reference sites. In contrast, a significant positive
relationship was found in those specimens from loca-
tions with high-water current, structural refugee,
submerged riparian coverage and low percentage of
detritus. These findings corroborate previous knowl-
edge of the habitat preferences of B. haasi (i.e.
rheophilous) and confirm that hydromorphology
strongly conditions the distribution of fish in a Medi-
terranean river (Sostoa, 1990; Doadrio, 2001). Though
the CF may serve as an initial screening biomarker in
fish health assessment, it mainly indicates the nutritional
status of fish and prey availability (Van der Oost et al.,
2003). In this regard, macroinvertebrate assemblages
change along a pollution gradient, but potential prey for
B. haasi are expected to be available in all conditions
due to its generalist feeding habits (e.g. chironomidae,
ephemeroptera, etc.) (Sostoa, 1990; Murria, 2003).
Fish population characteristics and biotic indices
Diagnostic discordances were detected between the
three approaches (i.e. water physicochemical param-
eters, general biotic indices and fish features) used to
determine the ecological status of Vallvidrera creek
and also the health status of this fish population.
Furthermore, the diagnoses of the two biotic indices
IBICAT and QBR were not consistent with each other.
This may be explained because QBR is an index of
riparian coverage (QBR in Munne et al., 1998) and
IBICAT a fish index, and fish do not directly depend
on the quality of riparian coverage, as long as they
have somewhere that acts as a refuge.
QBR scores drop at the most polluted sites.
Locations near built-up areas have the most altered
riparian coverage, but in addition the stream received
the inputs of urban collectors at these sampling sites.
The low score was due to the replacement of riparian
coverage by cultivated lands or the presence of
opportunists and allochthonous species that down-
weighted the index (i.e. the giant cane, Arundo donax)
(Munne et al., 1998; Murria, 2003).
IBICAT scoring did not vary between sampling
sites despite the impairments detected in this fish
Hydrobiologia
123
population. The selection of metrics in IBIs is
generally conditioned in Mediterranean rivers by the
low-species richness and the covariation of natural and
anthropogenic perturbations (Oberdorff et al., 2002;
EU FAME, 2004; Noble et al., 2007; Pont et al., 2007;
Benejam et al., 2008). In this case, the poor sensitivity
could be due to the selection of the reference water
body for this small river and the monospecific
condition of this fish community (Benejam et al.,
2008). This fact may explain that IBIs only reflect
strong changes at the population or community scale
(i.e. native species extirpation, big fluctuations in fish
density), without taking into account the subtle
changes that appeared in the population before (i.e.
biases in population structure). In addition, the metric
of fish density and the thresholds established may be
quite controversial since the capacity of these ecosys-
tems is not calculated, and natural big fluctuations are
reported in fish populations from Mediterranean rivers
(Magalhaes et al., 2007). Our findings argue for the
need to incorporate new metrics (e.g. fish population
structure) in the present version of IBICAT, in order to
detect the impact on fish populations that are over-
looked by the traditional biomonitoring protocols in
Mediterranean streams (Benejam et al., 2008; Mace-
da-Veiga et al., 2010b). The use of biomarkers using
non-invasive sample techniques (i.e. blood parame-
ters) may be a complementary tool in order to detect
sub-lethal effects on these fish populations, in which
low-fish richness difficult the development of power-
ful sensitive biotic indices.
Conclusions
This study revealed that the chronic exposure to the
effluents from a sewage treatment is responsible for
the deleterious effects detected on a population of
B. haasi in Vallvidrera creek. Skewed fish population
structure and low-fish density were the main changes
reported on this fish population. This study also
confirms the preference of B. haasi for fast-flowing
reaches. However, these effects were unnoticed when
IBICAT was applied as diagnostic tool. The selection
of the reference community for this small creek
together with its monospecific condition may explain
this finding. However, we suggest the incorporation of
new metrics (i.e. fish age or size structure) in the
present version of IBICAT to increase its diagnostic
power. Additionally, we recommend the use of
complementary non-lethal diagnostic tools (i.e. bio-
markers). Resource managers should be able to detect
subtle changes in fish communities before the effects
of anthropogenic impacts become evident at popula-
tion scale. Our findings also argue the need to change
water management policies in order to improve the
water quality of the effluents from sewage treatment
plants and to guarantee the dilution capacity of
Mediterranean rivers.
Acknowledgments We thank the team that performed fish
surveys and Robin Rycroft for revising the English. We also
thank the Collserola Natural Park and the Catalan Water Agency
for financial support and Dr Alex Richter for his ideas in the
previous draft of this study.
References
Adams, M. S., W. D. Crumby, M. S. Greeley Jr, L. R. Shugart &
C. F. Saylor, 1992. Response of fish populations and
communities to pulp mill effluents: a holistic assessment.
Ecotoxicology and Environmental Safety 24: 347–360.
Aparicio, E. & A. de Sostoa, 1998. Reproduction and growth of
B. haasi in a small stream in the N. E. of the Iberian pen-
insula. Archiv fur Hydrobiologie 142(1): 95–110.
Aparicio, E. & A. de Sostoa, 1999. Pattern of movements of
adult Barbus haasi in a small Mediterranean stream.
Journal of Fish Biology 55: 1086–1095.
Beche, L. A., P. G. Connors, V. H. Resh & A. M. Merenlender,
2009. Resilience of fishes and invertebrates to prolonged
drought in two California streams. Ecography 32: 778–788.
Benejam, L. E. Aparicio, M. J. Vargas, A. Vila-Gispert & E.
Garcıa-Berthou, 2008. Assessing fish metrics and biotic
indices in a Mediterranean stream: effects of uncertain
native status of fish. Hydrobiologia 603: 197–210.
Boix, D., E. Garcia-Berthou, S. Gascon, L. Benejam, E. Tornes,
J. Sala, J. Benito, A. A. Munne, C. Sola & S. Sabater, 2010.
Response of community structure to sustained drought in
Mediterranean rivers. Journal of Hydrology 383: 135–146.
Cambra, J., I. Alvaro, A. Gomez-Bolea & N. L. Hladun, 2000.
Estudis algologics a la Serra de Collserola i qualitat de
l’aigua de la riera de Vallvidrera (Barcelona). I Jornades
sobre la Recerca en els sistemes naturals de Collserola:
aplicacions a la gestio del parc. Conservation 83: 269–278.
CEN. 2003. Water Quality—Sampling of Fish with Electricity.
European standard—EN 14011:2003. European Commit-
tee for Standardization, Brussels, p. 18.
Clavero, M., Q. Pou-Rovira & L. Zamora, 2009. Biology and
habitat use of three-spined stickleback (Gasterosteusaculeatus) in intermittent Mediterranean streams. Ecology
of Freshwater Fish 18(4): 550–559.
Crawley, M. J., 2002. Statistical Computing: an Introduction to
Data Analysis Using S-Plus. Wiley, New York.
Crivelli, A. J., 2005. Barbus haasi. In: IUCN 2006. 2006 IUCN
Red List of Threatened Species.
Hydrobiologia
123
Damasio, J. B., C. Barata, A. Munne, A. Ginebreda, H. Guasch,
S. Sabater, J. Caixach & C. Porte, 2007. Comparing the
response of biochemical indicators (biomarkers) and bio-
logical ındices to diagnose the ecological impact of an oil
spillage in a Mediterranean river (NE Catalunya, Spain).
Chemosphere 66: 1206–1216.
Damasio, J. B., M. Fernandez-Sanjuan, J. Sanchez-Avila, S.
Lacorte, N. Prat, M. Rieradevall, A. M. V. M. Soares &
C. Barata, 2011. Multi-chemical responses of benthic
macroinvertebrate species as a complementary tool to
diagnose the cause of community impairment in polluted
rivers. Water Researh 45: 3599–3613.
Devlin, R. H. & Y. Nagahama, 2002. Y. Sex determination and
sex differentiation in fish: an overview of genetic, physi-
ological, and environmental influences. Aquaculture 208:
191–364.
Di Giulio, R. T. & D. E. Hinton, 2008. The Toxicology of
Fishes. CRC Press, Boca Raton: 1072 pp.
Doadrio, I., 2001. Atlas y Libro Rojo de los Peces Continentales
de Espana. Direccion General de la Natureza—Museo
Nacional de Ciencias Naturales, Madrid: 364 pp.
EU FAME, 2004. Development, Evaluation and Implementa-
tion of a Standardized Fish-based Assessment Method for
the Ecological Status of European Rivers. A Contribution
to the Water Framework Directive. European Project:
EVK1-CT-2001-00094.
Fernandez-Delgado, C. & M. Herrera, 1995. Age structure,
growth and reproduction of Rutilus lemmingii in an inter-
mittent stream of the Guadalquivir river basin, southern
Spain. Hydrobiologia 299(3): 207–213.
Gafny, S., M. Goren & A. Gasith, 2000. Habitat condition and
fish assemblage structure in a coastal mediterranean stream
(Yarqon, Israel) receiving domestic effluent. Hydrobiolo-
gia 422(423): 319–330.
Gasith, A. & V. H. Resh, 1999. Streams in Mediterranean cli-
mate regions: abiotic influences and biotic responses to
predictable seasonal events. Annual Review of Ecology
and Systematics 30: 51–81.
Grossman, G. D. & A. de Sostoa, 1994. Microhabitat use by
fishes in the lower Rio Matarrana, Spain: 1984-1987.
Ecology of Freshwater Fish 3: 123–136.
Grossman, G. D. & M. C. Freeman, 1987. Microhabitat
use in a stream fish assemblage. Journal of Zoology 212:
151–176.
Herrera, M., 1991. Estrategias en los ciclos de vida de una co-
munidad de ciprınidos en un arroyo de primer orden de la
cuenca del rıo Guadalquivir. Doctoral Thesis, Universidad
de Cordoba.
Hooke, J. M., 2006. Human impacts on fluvial systems in the
Mediterranean region. Geomorphology 79: 311–335.
Karr, J. R., 1991. Biological integrity: a long neglected aspect of
water resource management. Ecological Applications 1:
66–84.
Legendre, P. & L. Legendre, 1998. Numerical Ecology, 2nd ed.
Elsevier, Amsterdam.
Maceda-Veiga, A. & A. de Sostoa, 2011. Observational evi-
dence of the sensitivity of some fish species to environ-
mental stressors in Mediterranean rivers. Ecological
Indicators 11: 311–317.
Maceda-Veiga, A., H. Salvado, D. Vinyoles & A. Sostoa, 2009.
Ichthyophthirius multifiliis outbreaks in red-tailed barbel
Barbus haasi in a Mediterranean stream during drought.
Journal of Aquatic Animal Health 21: 189–194.
Maceda-Veiga, A., M. Monleon-Getino, N. Caiola, F. Casals &
A. de Sostoa, 2010a. Changes in fish assemblages in
catchments in north-eastern Spain: biodiversity, conser-
vation status and introduced species. Freshwater Biology
55(8): 1734–1746.
Maceda-Veiga, A., M. Monroy, G. Viscor & A. de Sostoa,
2010b. Changes in non-specific biomarkers in the Medi-
terranean barbel (Barbus meridionalis) exposed to sewage
effluents in a Mediterranean stream (Catalonia, NE Spain).
Aquatic Toxicology 100(3): 229–237.
Magalhaes, M. F., P. Beja, C. Canas & M. J. Collares-Pereira,
2002. Functional heterogeneity of dry-season refugia
across a Mediterranean catchment: the role of habitat and
predation. Freshwater Biology 47: 1919–1934.
Magalhaes, M. F., P. Beja, I. Schlosser & J. Collares-Pereira,
2007. Effects of multiyear droughts on fish assemblages of
seasonally drying Mediterranean streams. Freshwater
Biology 52: 1494–1510.
Magalhaes, M. F., C. E. Ramalho & M. J. Collares-Pereira,
2008. Assessing biotic integrity in a Mediterranean
watershed: development and evaluation of a fish-based
index. Fisheries Management and Ecology 15: 273–289.
Mariz, D., 2011. Investigacion sobre la eliminacion de olores en
depuradoras. PhD thesis, Universidad Politecnica de
Madrid, Madrid.
Munne, A., M. A. Sola & N. Prat, 1998. QBR: Un ındice rapido
para la evaluacion de la calidad de los ecosistemas de ri-
bera. Tecnologıa del agua 175: 20–37.
Murria, C., 2003. La qualitat ecologica de la riera de Vallvidrera
(Collserola). [The ecological quality of the Vallvidrera
creek]. Diploma d’Estudis Avancats, Universitat de Bar-
celona, Barcelona, Spain.
Noble, R. A. A., I. G. Cowx, D. Goffaux & P. Kestemont, 2007.
Assessing the health of European rivers using functional
ecological guilds of fish communities: standardizing spe-
cies classification and approaches to metric selection.
Fisheries Management and Ecology 14: 381–392.
Nogouchi, K., W. L. Wallace Hui, Y. R. Gel, J. L. Gastwirth &
W. Miao, 2009. lawstat: An R package for biostatistics,
public policy, and law. R package version 2.3.
Oberdorff, T., D. Pont, B. Hugueny & J. Porcher, 2002.
Development and validation of a fish-based index for the
assessment of ‘river health’ in France. Freshwater Biology
47: 1720–1734.
Oliva-Paterna, F. J., A. Vila-Gispert & M. Torralva, 2003.
Condition of Barbus sclateri from semi-arid aquatic sys-
tems: effects of habitat quality disturbances. Journal of Fish
Biology 63: 699–709.
Pont, D., B. Hugueny & C. Rogers, 2007. Development of a fish-
based index for the assessment of river health in Europe:
the European Fish Index. Fisheries Management and
Ecology 14: 427–439.
Pires, A. M., I. G. Cowx & M. M. Coelho, 1999. Seasonal
changes in fish community structure of intermittent streams
in the middle reaches of the Guadiana basin, Portugal.
Journal of Fish Biology 54: 235–249.
Pires, D. F., A. M. Pires, M. J. Collares-Pereira & F. Magalhaes,
2010. Variation in fish assemblages across dry-season
pools in a Mediterranean stream: effects of pool
Hydrobiologia
123
morphology, physicochemical factors and spatial context.
Ecology of Freshwater Fish 19: 74–86.
Prat, N. & J. V. Ward, 1994. The Tamed River. In Margalef, R.
(ed.), Limnology Now: A Paradigm of Planetary Problems.
Elsevier Science, London: 219–236.
Prat, N. & A. Munne, 2000. Water use and quality and stream
flow in a Mediterranean stream. Water Research 34:
3876–3881.
R Development Core Team, 2010. R: A Language and Envi-
ronment for Statistical Computing. R Foundation for Sta-
tistical Computing, Vienna.
Revelle, W., 2010. Psych: Procedures for Personality and Psy-
chological Research. Northwestern University, Evanston,
R package version 1.0-92. http://personality-project.org/r/
psych.manual.pdf, 1.0-90 edition.
Schiemer, F., 2000. Fish as indicators for the assessment of the
ecological integrity of large rivers. Hydrobiologia 422(423):
271–278.
Sedell, J. R., G. H. Reeves, F. R. Hauer, J. A. Stanford &
C. P. Hawkins, 1990. Role of refugia in recovery from
disturbances: modern fragmented and disconnected river
systems. Environmental Management 14(5): 711–724.
Soriguer, M. C., R. Bravo, C. Vallespin, C. Gomez-Cama & J.
A. Hernando, 2000. Reproductive strategies of two species
of cyprinids in a stream with Mediterranean regimen (SW
Spain). Archiv fur Hydrobiologie 148(1): 119–134.
Sostoa, A. (ed.), 1990. Historia Natural dels Paısos Catalans.
Peixos. Enciclopedia Catalana, Barcelona. 487.
Sostoa, A., N. Caiola, D. Vinyoles, S. Sanchez & C. Franch,
2003. Development of a Biotic Integrity Index (IBICAT)
Based on the Use of Fish as Indicators of the Environ-
mental Quality of the Rivers of Catalonia. Report to the
Catalan Water Agency, Barcelona (in Catalan).
Van der Oost, R., S. C. C. Lopes, H. Komen, K. Satumalay, R.
van den Bos, H. Heida, et al., 1998. Assessment of envi-
ronmental quality and inland water pollution using bio-
marker responses in caged carp (Cyprinius carpio): Use of
a bioactivation:detoxication ratio as biotransformation
index (BTI). Marine Environmental Pollution 46: 315–319.
Van der Oost, R., J. Beyer & N. P. E. Vermeulen, 2003. Fish
bioaccumulation and biomarkers in environmental risk
assessment: a review. Environmental Toxicology and
Pharmacology 13: 57–149.
Verdiell, D., F. J. Oliva-Paterna & M. Torralva, 2007. Fish
condition of Barbus haasi Mertens, 1925 in the Chico
River (Castellon, northern east Iberian Peninsula). Duga-
stella 4: 43–47.
Vila-Gispert, A., E. Garcıa-Berthou & R. Moreno-Amich, 2002.
Fish zonation in a Mediterranean stream: effects of human
disturbances. Aquatic Science 64: 163–170.
Vinyoles, D., A. De Sostoa, C. Franch, A. Maceda-Veiga, F.
Casals & N. Caiola, 2010. Life-history traits of the stone
loach Barbatula barbatula. Journal of Fish Biology 77(1):
20–32.
Welch, E. B. & T. Lindell, 1992. Ecological Effects of Waste-
water, Applied Limnology and Pollutant Effects. Spon,
London.
Hydrobiologia
123
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