Benthic community structure and biomarker responses of the clam
Scrobicularia plana in a shallow tidal creek affected by fish farm effluents
(Rio San Pedro, SW Spain)
Claudio Silvaa,b,d,⁎, Mattia Mattiolic, Elena Fabbric, Eleuterio Yáñezd,
T. Angel DelVallsa, M. Laura Martín-Díaza,b
aUNITWIN/UNESCO/WiCoP, Physical Chemical Department, University of Cádiz, Campus de Excelencia Internacional del Mar (CEIMAR),
Polígono Río San Pedro s/n, 11510 Puerto Real, Cádiz, Spain
bAndalusian Center of Marine Science and Technology (CACYTMAR), University of Cádiz, CEIMAR, Polígono Río San Pedro s/n, 11510 Puerto Real, Cádiz, Spain
cInterdepartment Centre for Research in Environmental Sciences (CIRSA), University of Bologna, Via S. Alberto 163, 48123 Ravenna, Italy
dSchool of Marine Science, Pontificia Universidad Católica de Valparaíso, Avda. Altamirano 1480, Casilla 1020, Valparaíso, Chile
a b s t r a c ta r t i c l ei n f o
Received 31 October 2011
Accepted 17 June 2012
Available online 13 July 2012
Marine fish farm impact
Rio San Pedro creek
The effects of solid organic wastes from a marine fish farm on sediments were tested using benthic commu-
nity as ecological indicators and biomarkers in native clam (Scrobicularia plana) as biochemical indicators.
The benthic fauna and clam samples were collected in the intertidal sediment in October 2010 from five
sites of the Rio San Pedro (RSP) creek, following a gradient of contamination from the aquaculture effluent
to the control site. Numbers of species, abundance, richness and Shannon diversity were the biodiversity
indicators measured in benthic fauna. Morphological and reproduction status of clams were assessed using
the condition factor and gonado-somatic index, respectively. Phase I and Phase II detoxification enzymatic
activities (ethoxyresorufin O-deethylase (EROD), glutathione S-transferase (GST)), antioxidant enzymatic
activities (glutathione peroxidase (GPX), glutathione reductase (GR)) and oxidative stress parameters
(Lipid Peroxidation (LPO) and DNA strand breaks) were measured in clams' digestive gland tissues. In paral-
lel, temperature and salinity in the adjacent water, redox potential, pH and organic matter in sediment, and
dissolved oxygen in the interstitial water were measured. The results suggested that RSP showed a spatial
gradient characterised by hypoxia/anoxia, reduced potential, acidic conditions and high organic enrichment
in sediments at the most contaminated sites. Significant (pb0.05) decrease of biodiversity indicators were
observed in the areas impacted by the aquaculture discharges. Biomarkers did not show a clear pattern
and of all biochemical responses tested, GPX, DNA damage and LPO were the most sensitive ones and showed
significant (pb0.05) increase in the polluted sites. Benthic biodiversity indicators were significantly (pb0.05)
positively correlated with pH, redox potential and dissolved oxygen and negatively correlated with organic
matter. On the contrary, antioxidant enzymatic responses (GPX) and oxidative stress parameters were signif-
icantly (pb0.05) negatively correlated with those physico-chemical parameters. It has been demonstrated
that effluents from fish aquaculture activities in Río San Pedro creek may produce an alteration of physico-
chemical characteristics of seabed and induce oxidative stress and DNA damage in soft-sediment species
which may lead to changes of the benthic population structure and health status of the exposed organisms.
© 2012 Elsevier Ltd. All rights reserved.
Aquaculture is one of the fastest growing food-producing sectors,
supplying approximately 47% of the world's fish food (FAO, 2009).
However, this strong expansion in the fish aquaculture industry has
brought significant marine environmental impacts in the littoral eco-
systems, such as sediment organic enrichment and eutrophication
(Holmer et al., 2005; Kalantzi and Karakassis, 2006), chemical pollution
from pharmaceuticals, organics, bactericides and metals (Antunes and
Gil, 2004; Cabello, 2006; Sapkota et al., 2008), and changes in the biodi-
versity and community structure of benthic fauna (Tomassetti et al.,
2009; Vezzulli et al., 2008).
Rio San Pedro (RSP) is a salt marsh creek situated in the southwest
of Spain and is a highly productive area where aquaculture and clam
fishing activities have traditionally taken place. It is also important
from an ecological point of view as a result of its qualification as a
Environment International 47 (2012) 86–98
⁎ Corresponding author at: UNITWIN/UNESCO/WiCoP, Physical Chemical Department,
University of Cádiz, Campus de Excelencia Internacional del Mar (CEIMAR), Polígono
Río San Pedro s/n, 11510 Puerto Real, Cádiz, Spain. Tel.: +34 956016794; fax: +34
E-mail address: email@example.com (C. Silva).
0160-4120/$ – see front matter © 2012 Elsevier Ltd. All rights reserved.
Contents lists available at SciVerse ScienceDirect
journal homepage: www.elsevier.com/locate/envint
Natural Park in 1996. Despite this qualification, untreated aquaculture
waste waters from a fish farm effluent in the upper part of the creek dis-
charge directly to the coastal waters and constitute a major factor of pol-
lution. The land-based marine fish farm is devoted to the semi-intensive
culture of gilthead sea bream (Sparus aurata) and European sea bass
contamination is produced by the waste in dissolved and particulate
siderable amount of uneaten food and faeces settles out as sediment
(Karakassis et al., 2002; Papageorgiou et al., 2009). The area affected
by the organic enrichments of the fish farm effluent in the RSP is
characterised by a low pH and increased levels of nutrients, organic
matter and suspended solids (De la Paz et al., 2008; Tovar et al.,
2000a). The organic enrichment may change the physical and chemical
composition of the sediments (Karakassis et al., 2002), affecting the
structure of the benthic communities and health statuses of the biota
(Solan et al., 2004).
In terms of environmental/ecological impact and risk assessment,
the monitoring of pollution (contamination that causes adverse biolog-
ical effects in the natural environment) effects requires the develop-
ment and application of a variety of robust and rapid tools (Chapman,
tion, determination of benthic community structure and assessment of
toxicity using biomarkers was proven to be useful for establishing
evidence of exposure to pollutants and damage to thehealth of sentinel
organisms (Chapman, 1990; Moore et al., 2004; Van der Oost et al.,
2003). Ecological monitoring using benthic structure analysis will
provide information on changes in species composition, abundance
and diversity, all of which may be indicative of the effects of pollution
in communities (Chapman, 1990, 1996; Chapman et al., 1987; DelValls
et al., 1998; Long and Chapman, 1985). Biological effect monitoring
using biomarkers of exposure provides information on early adverse al-
terations that occurinorganismsaftercontaminants havebeentakenup
either via respiratory surfaces or via ingestion. This monitoring may, in
some cases, provide early warning of potential future organism-level
effects (Chapman, 2007; Galloway et al., 2004; Martín-Díaz et al.,
2004). However, the use of biomarkers in monitoring programmes is
still limited due to: a) temporal and spatial variability of biochemical
responses, and b) natural abiotic and biotic factors, called confounding
factors, which include for example temperature, salinity, season, sex,
age, condition factor and reproductive status (Fossi Tankoua et al.,
2011; Hagger et al., 2006; Handy et al., 2003). In recent years, an in-
creased number of scientific papers have been published with respect
to the application of benthic community structure (Edgar et al., 2005;
Papageorgiou et al., 2009; Tomassetti and Porrello, 2005) and bio-
markers (Bocchetti et al., 2008; Carballeira et al., 2011; Cravo et al.,
2012; Solé et al., 2009) as tools for the assessment and biomonitoring
of pollution effects in areas affected by fish farm effluents. However, to
our knowledge no information is available concerning the application
of benthic faunal community and biomarker indicators to assess the im-
pact of semi-intensive earthen fish ponds and there is a lack of informa-
tion about the links between the environmental behaviour of the
ical organisations (individual and community health status) in the area
of the RSP creek affected by the aquaculture effluents discharges.
This work presents an integrated biomonitoring approach to study-
ing the effects of marine fish farm effluent on sediment characteristics,
the benthic fauna community structure and biochemical biomarkers
from field-collected native clams Scrobicularia plana. The aims were to
evaluate the suitability of benthic and biomarker indicators for sediment
quality assessment and to study the cause-and-effect relationships be-
tween physico-chemicals, the benthic structure indicators and biomark-
er responses, all of which improve biomonitoring and provide an early
warning detection system for environmental quality and ecological risk
assessments in coastal marine ecosystems (Bolognesi et al., 2004;
DelValls et al., 1998; Martín-Díaz et al., 2008). The battery of biomarkers
used in this study are recommended for pollution monitoring studies
(Martín-Díaz et al., 2004) and included detoxification enzymatic activi-
ties (ethoxyresorufin O-deethylase—EROD, glutathione S-transferase—
GST), antioxidant enzymatic activities (glutathione peroxidase—GPX,
glutathione reductase—GR) and oxidative stress parameters (Lipid
Peroxidation—LPO and DNA strand breaks) that were measured in
clam digestive gland tissues.
2.1. Study area
The research work was carried out in the RSP (36°23′–37′N, 6°8′–
15′W) (Fig. 1). The RSP is a shallow tidal creek within the salt marsh
area of the Bay of Cadiz (SW Spain). The area was selected because a
fish farm facility is located in the upper part of the creek. It used to be
a tributary of the Guadalete River until it was artificially blocked at
12km from the mouth, so currently the only freshwater input into
the creek comes from precipitation and land drainage. The RSP
creek has complementary and in some cases conflicting uses, includ-
ing recreational shipping, fishing boat anchorages, semi-intensive
marine fish aquaculture, urban centres, touristic beaches and natural
protected areas (Fig. 1). A fish farm in earthen ponds is located in the
upper part of the creek, producing 550tonsyear−1of gilthead sea
Fig. 1. Location of study sites in the Río San Pedro creek. The map shows the main land
uses. Points with numbers in the marine coastal area correspond to sampling locations.
C. Silva et al. / Environment International 47 (2012) 86–98
bream Sparus aurata and European sea bass Dicentrarchus labrax
(Empresa Pública Desarrollo Agrario y Pesquero, Consejería de
Agricultura y Pesca). The fish are fed with industrial feed. Point source
discharges of effluents from the fish farm are discharged into the
creek, and constitute a major factor of pollution and deterioration of
environmental quality in the RSP creek.
The RSP is subject to a semi-diurnal tidal regime with the height of
the tidal column varying from 3.5m at spring tide to 0.5m at neap
tide. The creek has a maximum depth of 4–5m and water input
comes from tides from the Bay of Cadiz making water renewal in the
upper part (aquaculture zone) to be very poor. In the upper part of
culture, thus the only inputs received by water are fish farm effluents,
water renovation from the Bay and rainwater. Any non-natural alter-
ation of the sediment must be related to the aquaculture activities.
studied the longitudinal distribution of various physico-chemical prop-
erties in the RSP, and suggested that the low pH values in the fish farm
effluent are due to the high ammonium concentration and to the acidic
character of the faeces and the industrial fish feed. The seasonal pattern
of nutrients and other physico-chemical parametersinthe RSP isdiffer-
ent from the known seasonal pattern for coastal waters (Tovar et al.,
2000a), and it follows the growth rate curve of the fish cultivated on
the farm (Tovar et al., 2000b). This temporal pattern is related to the
high seasonality in the material inputs to the RSP (mainly nutrients, or-
ganic matter and suspended particulate matter) originating from the
fish farm effluent discharges, according to the results obtained by
Tovar et al. (2000b) who studied the temporal variability in the out-
flows from the fish farm over a 2-year period. An increase was found
in the fish production cycle with temperature, with maxima in summer
and autumn, and consequently the highest discharges of nutrients, dis-
solved organic carbon and matriculate material to the RSP creek. Chlo-
rophyll a values follow this trend in response to high nutrient inputs,
mainly nitrites and nitrates, and to the favourable conditions caused
by residence time (De la Paz et al., 2008). Tovar et al. (2000a) identify
the existence of two different zones within the inlet according to
water quality: the first one with a length of about 8km is closer to the
mouth and there are no fish farm and have a good renovation of the
water controlled by the tides and the second section is affected by the
semi-intensive fish culture and its effluents are considered the main
source of nutrients, organic matter and suspended solids received by
2.2. Sampling collection
Samplings were carried out in autumn (October) 2010 during low
tide. This season was selected since it corresponds to the period of
maximum fish production and discharges to the creek (De la Paz et
al., 2008; Tovar et al., 2000b), in order to reflect the highest potential
impact of the farms. A total of 5 sampling sites (RSP1, RSP2, RSP3,
RSP4 and RSP5) were established at 0.05, 0.3, 2.5, 4.3 and 9.5km, re-
spectively, following a gradient of contamination from the aquacul-
ture effluent to the reference site (RSP5) located in the lower part
of the creek in the area not affected by the fish farm (Fig. 1). RSP5
was assumed as a control site due to being located in an area of
good water renovation (Tovar et al., 2000a), which has been previ-
ously used as a reference control zone in pollution biomarker studies
(Pérez et al., 2004; Solé et al., 2009).
Water samples were collected at 1m depth to measure tempera-
ture (T) and salinity (SA) directly in field. Three replicate samples
were taken at each site.
Sediment samples were taken at each sampling site using a
0.025m2Van Veen grab. Six replicate (independent Van Veen grab)
samples were taken at each site, three for benthic faunal analysis
and three for chemical analysis. After collection, sediments samples
were maintained at 4°C until their analysis in the laboratory (no
more than 4days).
In order to minimise the confounding factors that may influence
biomarker responses, a total of 30 native S. plana clams of similar
size (mean shell length 3.54±0.2cm) were collected by hand on
the same day at each sampling site in the lower intertidal sediment
during low tides. According to Sola (1997) the individuals of 3.3cm
represent adult clams of more than three years due to S. plana
reproduced for the first time when they reached the age of two
years. Following collection, clams were immediately placed in plastic
bags and kept at 4°C in thermally insulated 10-L boxes for transport
to the laboratory for morphometrics, reproductive and biochemical
2.3. Physico-chemical measurements
The T was registered using a VWR CO3000H electrode (ref.
663-0143) with an accuracy of ±0.1°C. The SA was measured with
a VWR CO3000H electrode with an accuracy of ±0.01psu (practical
salinity units), which was previously calibrated with a control stan-
dard solution of 0.01molL−1of KCl. Redox potential (Eh) and pH
were measured immediately after collection in sediment samples at
a depth of 3cm. The Eh was measured (accuracy of ±0.1mV) with a
Redox Buffer standard (Cat. No. 1412060) of known mV value (475mV
at 25°C). The pH was measured with a VWR ph1000H electrode (ref.
662-1151) previously calibrated with two buffers (pH=7 and 10) and
tial water of the sediments was determined using a VWR OX4000H
electrode (ref. 664-0038) with an accuracy of 0.001mgL−1. Sediments
samples were stored and dried at room temperature until they were
processes for measured the total organic matter (TOM) content which
was obtained through calcination (loss of ignition) in a muffle furnace
at 450°C for 5h with an accuracy of ±0.1%.
2.4. Benthic community analysis
In the laboratory, sediment samples were sieved using a 0.5mm
mesh size sieve and retained fauna were fixed in 10% buffered forma-
lin. After 3days the samples were preserved in 70% ethanol until the
benthic community analysis was performed. Benthic fauna samples
were sorted and organisms were counted and identified to the lowest
taxonomic level under a stereo microscope, using available species
keys (Ocaña-Martín et al., 2000; Riedl, 1986). The biological material
collected included both macrofauna and meiofauna (nematodes). The
total species (S) value was calculated as the total number of species
collected at each sampling site. The mean abundance (number of in-
dividuals per 0.025m−2; N), Margalef species richness (d; Margalef,
1963, 1968) and Shannon diversity (H′; Pielou, 1974) were also
calculated for each sampling site. Univariate measures of benthic faunal
data were performed with the computer software package PRIMER
6.1.6 (Plymouth Routines in Multivariate Ecological Research) (Clarke
and Warwick, 2001).
2.5. S. plana tissues processing, morphometrics and reproductive status
The clams were brought to the laboratory and each animal was
measured (shell length in mm) and weighed (g wet weights of
clams and soft tissues) on ice to determine its condition factor (CF)
(g clam weight/mm shell length). Gonadal and digestive gland (DG)
tissues were dissected on ice and weighed. The gonado-somatic
index (GSI) was determined by gonad wet weight/soft tissues wet
weight. Five pools of 6 clam DGs per sampling site were prepared,
frozen and stored at −80°C until biomarker analysis (no more than
1month). Only DGs were used for biomarker analysis because it is
the most sensitive tissue to sediment pollution as was supported by
C. Silva et al. / Environment International 47 (2012) 86–98
many research works in which biomarker determinations in clams
were conducted directly in this tissue (Romero-Ruiz et al., 2008;
Solé et al., 2009; Zhang et al., 2010).
2.6. Biochemical biomarker analysis in S. plana
Clam DG tissues were homogenised (Ultra-turrax homogeniser)
with a buffer containing dithiotheritol (1:3 weight:volume relation).
Homogenates were used to analyse the total protein (TP), LPO and
DNA strand breaks. The remainder was centrifuged (15,000g; 4°C;
20min) to obtain the S15 fraction, which was employed for TP,
EROD, GST, GPX and GR activity analyses.
TP was determined in the homogenised fraction and in the S15
fraction following an adaptation of the Bradford (1976) methodology.
Briefly, samples were incubated with diluted Bio-Rad protein assay
reagent (Bio-Rad Laboratories GmbH Cat. No. 5000-0006). Absor-
bance at 595nm was measured with a spectrophotometer (Model
Infinite M200 PRO, TECAN, Austria) and the TP was quantified using
the Bradford protein assay of the SOFTmax PRO software package
(Molecular Devices, USA). TP concentrationwasexpressed asmgmL−1.
EROD is a Phase I biotransformation enzyme in the xenobiotics me-
tabolism. Phase I of metabolismconsists of theoxidation,reduction and
activity was measured using the adaptation of the Gagné and Blaise
(1993) method previously conducted by Martín-Díaz et al. (2008).
The rate of resorufin production through EROD mediated deethylation
of the substrate 7-ethoxyresorufin was determined fluorometrically
using 516nm (excitation) and 600nm (emission) filters. The reaction
was initiated by the addition of NADPH and continued for 60min. A
results into pmol of resorufinmin−1mg−1of total protein.
GST is a Phase II biotransformation enzyme that catalyses electro-
philic compounds and Phase I product conjugation with reduced glu-
tathione (GSH) (Van der Oost et al., 2003). Additionally, GST presents
antioxidant action (Pereira et al., 2009). The methodology of
McFarland et al. (1999) was adapted according to Martín-Díaz et al.
(2008) for GST activity determination. The enzyme activity was
analysed using 1-chloro-2,4-dinitorbenzene (CDNB) and reduced glu-
tathione (GSH) as substrates and measured spectrophotometrically at
340nm every 5min for 30min. Results were expressed as nmol of
conjugate CDNBmin−1mg−1of total protein.
GPX is an antioxidant enzyme that deactivate reactive oxygen spe-
cies (ROS) generated during normal metabolism and by contaminants
(Livingstone, 2003). GPX catalyses H2O2transformation into water
with GSH oxidation. GPX activity was measured indirectly through
the GR-mediated reduction of oxidised glutathione (GSSG) according
to the methodology described by McFarland et al. (1999) and adapted
by Martín-Díaz et al. (2008). GPX activities were measured spectro-
photometrically at 340nm every 10s for 3min at 30°C, using a
1mM cumene hydroperoxide substrate. The analysis of GPX activity
measures the decrease in absorbance at 340nm resulting from the
consumption of NADPH by the GR reaction. Results were expressed
as pmol NADPHmin−1mg−1of total protein.
GR is an antioxidant enzymes that deactivate ROS and provides
GSH by catalysing the reduction of its oxidised form (Van der Oost
et al., 2003). GR activity was determined utilising an adaptation of
the McFarland et al. (1999) procedure previously conducted by
Martín-Díaz et al. (2008). GR activity, together with the co-factor
NADPH, catalyses the reduction of GSSG to GSH. The consumption
of NADPH produces a decrease in absorbance at 340nm, which is
directly proportional to the glutathione reductase activity in the sample.
Spectrophotometric measures were made at 340nm every 2min for
10min at 30°C. Results were expressed as nmol of NADPHmin−1mg−1
of total protein.
LPO and DNA strand breaks are early effects of oxidative stress
(Valavanidis et al., 2006) originated by the action of xenobiotics and/or
tation of the Thiobarbituric Acid Reactive Substances (TBARS) method
by Wills (1987) was used for LPO determination. Malondialdehyde
(Janero, 1990). MDA reacts with 2-thiobarbituric acid to produce tetra-
methoxipropane, which is measured spectrophotometrically at 540nm
to allow indirect determination of MDA. MDA concentration was calcu-
lated by means of a tetramethoxypropane standard curve. Results were
expressed as nmol of TBARSmg−1of total protein (homogenate).
DNA damage was assessed in accordance with the alkaline precipi-
tation assay (Olive, 1988). In the presence of detergents, genomic
DNA (associated with proteins) precipitates in alkali and leaves
protein-free and alkali-labile DNA strands in solution. These strands
are then measured using fluorescence. Homogenates were mixed with
a SDS solution and KCl was added. The mixture was heated to 60°C
and cooled to precipitate the genomic DNA linked to SDS. This mixture
was then centrifuged at 8000g for 5min at 4°C. Hoescht dye was
added to the supernatant. Fluorescence was measured using 360nm
(excitation) and 450nm (emission) filters. Salmon sperm genomic
DNA standardswereused forDNAcalibrationand theresults expressed
as μg of supernatant DNAmg−1of total protein (homogenate).
2.7. Statistical approach
Univariate analyses were based on the community descriptive pa-
rameters of the benthic fauna and the biomarker responses estimated,
which were calculated for each replicate sample and summarised for
each site. In order to determine whether there were significant differ-
ences among sites, a one-way analysis of variance (ANOVA) in each
benthic univariate measure and biomarker responses was used and
was followed by multiple comparisons of Dunett's tests. The signifi-
cance was set at pb0.05. Univariate statistical analyses were performed
using SPSS 15.0.
For multivariate analyses, the species abundance was transformed
ing (MDS), distance-based permutational multivariable analysis of
variance (PERMANOVA) and similarity percentage analysis (SIMPER).
Relative similarities in abundance of benthic fauna among sites were
graphically analysed using a MDS with Bray–Curtis similarity index
(Clarke and Gorley, 2006). The reliability of MDS representations of
the assemblage similarities was assessed with their stress values;
stressb0.2 was considered to be acceptable (Clarke and Warwick,
2001). The statistical significance of differences in the benthic fauna
abundance among sampling sites were analysed with PERMANOVA
based on Bray–Curtis similarity measures (Anderson et al., 2008).
PERMANOVA software was used for testing the simultaneous response
of fauna to one factor in a one-way ANOVA experimental design on the
basis of any distance measure, using permutation methods (Anderson,
2001). For the one-way case, an exact p-value was provided using
unrestricted permutation of raw data. In addition, a posteriori pairwise
tion distribution were available, asymptotical Monte Carlo p-values
(pMC) were used instead of permutational p-values (pPERM). A 1-factor
crossed design was used with sampling sites (5 levels: fixed) as factor
and three replicates. SIMPER analysis of abundances (cut-off percent-
ferences between site location and distance from the fish farm effluent
(Clarke and Warwick, 2001). All analyses of MDS, PERMANOVA and
SIMPER were performed using PRIMER 6.1.6.
Significant relationships between the distribution patterns of ben-
thic fauna abundance and physico-chemical features over the sampling
sites or gradientof contamination were exploredusingcanonical corre-
spondence analysis (CCA; Ter Braak and Verdonschot, 1995). Initial
mental variables (T, SA, Eh, pH, DO and TOM). The Monte Carlo permu-
C. Silva et al. / Environment International 47 (2012) 86–98
the significance of fauna–environmentrelationships. The level of signif-
was used for CCA analysis. Additionally, relationship between benthic
community structure and environmental variables were investigated
using the BIOENV program on PRIMER 6.1.6.
characteristics, benthic fauna and biomarker responses values, the
Pearson correlation coefficient was calculated using SPSS 15.0.
3.1. Physico-chemical characteristics
Summarised results for the physico-chemical characteristics of
adjacent water, sediment and interstitial water by sampling site are
shown in Table 1. No significant spatial differences were observed
in water SA and T in the sampling sites. A spatial gradient of contam-
ination was observed from the aquaculture effluent (RSP1) to the
control site (RSP5) in the muddy sediment chemical data. Lowers
levels of Eh (67.3mV), pH (7.138) and DO (0.034mgL−1) and higher
values of TOM (17.5%) were found in RSP1, which was the site closest
to the aquaculture effluent. On the contrary, in the control site (RSP5)
exhibited higher levels of Eh (124mV), pH (7.793) and DO
(1.557mgL−1), and lower levels of organic matter. All the sampling
sites were characterised by muddy bottom sediment.
3.2. Benthic community structure
A total of 1120individuals were identified in thebenthic faunaanal-
ysis at the five sampling sites, representing 26 species (9 mollusks, 5
polychaetes, 5 crustaceans, 4 nematodes and 3 oligochaetes). All the
univariate indices followed very similar trends at the different sites,
characterised by a spatial gradient with higher values in the control or
clean site and lower in the polluted sites close to the fish farm (Fig. 2).
S values ranged from 16 (control site RSP5) to 3 (RSP1), and a spatial
variability or gradient was observed with an increase in S from the
pollution effluent to the control site (Fig. 2a). Significant (pb0.05) dif-
ferences were observed between sampling sites (RSP1 and RSP2)
close to the effluent and the control site RSP5.
Significant (pb0.05) differences among the sampling sites were
also observed for N, with a spatial gradient characterised by greatest
values in the control site and lower values in the sites close to the
fish farm effluent (Fig. 2b). Significant (pb0.05) differences were
obtained between the sites adjacent (RSP1 and RSP2) to the fish
farm and the control for the d index (Fig. 2c). A spatial gradient was
found with higher values of species richness in the control site and
lower values in the upper part of the creek. The biodiversity index
H′ showed significant (pb0.05) differences between sampling sites
and followed similar spatial fluctuations (Fig. 2d).
The MDS ordination plot also supported the influence of the spa-
tial gradient of contamination from the aquaculture effluent (RSP1)
to the control site (RSP5) in benthic communities, showing that the
sampling sites were gathered into two groups at a 58% similarity
level (Fig. 3). The first group includes the control site RSP5 and
RSP4. In the second group, the sites correspond with the three sites
(RSP1, RSP2 and RSP3) nearest to the fish farm effluent (stress=
PERMANOVA revealed highly significant (pPERM=0.001) differ-
ences in benthic fauna composition among sampling sites (Table 2).
Pairwise tests showed that a highly significant (pMC=0.001) differ-
ence in fauna occurred between the control site (RSP5) and the site
nearest to the aquaculture effluent (RSP1). Significant (pMCb0.05) dif-
ference occurred among the other sites except for RSP2 and RSP3,
which have no statistical dissimilarities.
The SIMPER analysis showed how the benthic community contrib-
uted to site location or distance from the fish farm effluent, indicating
the main taxa abundance and contribution (Table 3). The results
showed levels of dissimilarity between the fauna abundance from
the sampling sites, with the higher dissimilarity between the control
sites RSP5 and RSP3, and RSP2 and RSP1 ranging between 54% and
65%. The control site was mainly characterised by mollusks (S. plana
and Hydrobia ulvae), oligochaetes and polychaetes (Paradoneis lyra,
Nereis diversicolor). The benthic community of the sites (RSP1, RSP2
and RSP3) located nearest to the fish farm effluent were characterised
by high abundance of the isopod Cyathura carinata and the polychaete
3.3. Morphological characteristics and reproductive status of S. plana
Individual and morphological characteristics of clam populations
collected in the sampling sites were examined by monitoring the CF
(Fig. 4). The CF was significantly increased (pb0.05) at sites nearest
(RSP1 and RSP2) to the fish farm effluent compared to the control
The reproductive status in clams was studied by tracking the GSI
(Fig. 4). Significant (pb0.05) differences between the control and
sampling sites were also observed for GSI, with greatest values in
RSP5 compared with the other sites.
3.4. Biomarker responses in S. plana
In Fig. 5, the activity of the different enzymes and oxidative stress
parameters measured in the DG of S. plana collected in the sampling
sites is shown. The EROD activity showed induction (RSP2 and
RSP3) and inhibition (RSP1 and RSP4) compared to the control site,
but the differences were not significant. GST activity showed an in-
crease compared to the control site in the DGs of clams from all sam-
pling sites except in RSP4, but the differences were not significant
(pb0.05). GPX activity was significantly (pb0.05) induced in the
DGs of S. plana from RSP1 (2.1-fold increase), RSP2 (2-fold increase)
and RSP3 (2.1-fold increase) compared with control site. GR activity
was significantly (pb0.05) induced in the DGs of clams from RSP3,
where the biomarker demonstrated an increase of 1.9-fold relative
to the biomarker results of the control site RSP5. Significant
(pb0.05) differences in LPO were measured in clams collected from
RSP1, RSP2 and RSP4, where the biomarker demonstrated increases
Average values and standard deviations of each physico-chemical variables of water (T and SA), sediment and interstitial water (DO) at the different sampling station in the Río San
Chemicals RSP1RSP2RSP3 RSP4 RSP5
C. Silva et al. / Environment International 47 (2012) 86–98
of 2.9-fold, 1.5-fold and 1.7-fold, respectively, compared with the
control site. DNA strand breaks showed significantly (pb0.05) higher
values in the DGs of clams collected in sediments from RSP1 (3-fold
increase) compared with the control site.
3.5. Relationships between physico-chemicals, biomarker responses and
The relation between site pattern in benthic fauna abundance and
environmental variables are presented by the CCA triplot (Fig. 6). The
Monte Carlo permutation test indicated a significant ordination dia-
gram (F ratio=2.83: pb0.05) in which the two first axes explained
90.88% of the total variance (67.35% in the first axis and 23.53% in
the second axis). Axis one was positively correlated with TOM con-
centration and negatively correlated with Eh, pH and DO. This indi-
cated an increase in organic matter, acidification, reduction and
anoxic conditions of sediments from left to right of the diagram
(Fig. 6), mainly evidenced at sampling sites (RSP1, RSP2 and RSP3)
nearest to the aquaculture effluent. The species coupling with such
conditions were C. carinata, Columbella rustica and Discus sp.
(Fig. 6). On the other extreme of the diagram was located the control
site RSP5 with Cerastoderma glaucum, Prionospio cirrifera, P. lyra and
S. plana as characteristic species of theprevailingenvironmental condi-
tions at this location. The second axis was positively correlated with T,
TOM and Eh, indicating an increase of these environmental variables
at RSP4 site. The fauna associated with these conditions were nema-
todes, and the annelids Pygospio elegans and Capitela capitata (Fig. 6).
If only one variable is considered in the BIOENV analysis, TOM was
found to present the best correlation coefficient (0.882) with benthic
fauna abundance (Table 4). This value slightly decreased (0.869)
when three variables (Eh, pH and TOM) were considered. A higher
number of variables (SA, T, Eh, pH and TOM) did not significantly
decrease (0.825) the value of the correlation coefficient.
No significative correlations were found between the clam's CF
and environmental measurements. The clam's GSI were significantly
(pb0.05) positively correlated with the distance to fish farm effluent
(r2=0.628), Eh (r2=0.531) and DO (r2=0.761), and negatively cor-
related with TOM (r2=−0.668). No significative correlations were
found between biomarker responses and the clam's CF and GSI. Few
significant correlations between biomarker responses and environ-
mental variables were observed. No significant correlations were
observed between physico-chemical data and EROD, GST and GR en-
zymatic activities. GPX (r2=−0.815), LPO (r2=−0.651) and DNA
strand breaks (r2=−0.519) biomarker responses were found to be
significantly (pb0.05) activated when decreasing the distance to the
source of discharges. GPX, LPO and DNA strand breaks were signifi-
cantly (pb0.05) activated due to the decrease of Eh, pH and DO, and
the increase of TOM. No significant correlations were found between
biomarker responses and the water SA and T. Regarding correlations
between benthic fauna measurements and biomarker responses, sig-
nificant (pb0.05) negative correlations were observed between biodi-
versity indicators (S, N, d and H′) and GPX, LPO, DNA strand breaks.
The benthic fauna and biomarker approach in this study permitted
the examination of relationships between environmental conditions,
benthic fauna population characteristics and biochemical responses
in clams. Higher contamination levels were observed at sites near
the aquaculture effluent and exposure to organic enrichment sedi-
ments from these sites caused adverse effects comparatively to the
control site, both at the population level in benthic fauna and cellular
level in clams. It has been demonstrated that discharges from fish
RSP1 RSP2 RSP3RSP4
Fig. 2. Mean values of the number of species (S), total abundance (N), richness (d) and Shannon diversity (H′) indices along the sampling stations in the Río San Pedro creek. Error
bars indicate standard deviations of mean based on 3 replicates at each site. RSP1–RSP5, Río San Pedro stations. Asterisks (*) indicate significant differences (pb0.05) compared with
the control site RSP5.
Fig. 3. Non metric multi-dimensional scaling (MDS) plot for fourth root transformed
total fauna abundance in sampling sites.
C. Silva et al. / Environment International 47 (2012) 86–98
SIMPER output indicating species contributing the most to the similarity within site location (distance to the fish farm effluent) for benthic fauna abundance. Shaded boxes: percent
similarity (bold) and the species that contributed to the similarity in each site. Non-shaded box: percent dissimilarity (bold) between sites location and the species that contributed
to the total dissimilarity (cut-off percentage: 85%).
RSP1 (0.05km) RSP2 (0.3 km) RSP3 (2.5 km) RSP4 (4.3 km)
RSP5 (9.5 km)
P ygospio elegans
Results of PERMANOVA based on Bray–Curtis dissimilarities (%) to test for differences in benthic fauna abundance at each sampling site (pMC-values are shown in parenthesis).
RSP1 vs. RSP2=33.28 (0.019)
RSP1 vs. RSP3=44.04 (0.009)
RSP1 vs. RSP4=59.54 (0.003)
RSP1 vs. RSP5=63.86 (0.001)
RSP2 vs. RSP3=31.82 (0.076)
RSP2 vs. RSP4=46.23 (0.013)
RSP2 vs. RSP5=54.24 (0.003)
RSP3 vs. RSP4=43.43 (0.016)
RSP3 vs. RSP5=50.35 (0.003)
RSP4 vs. RSP5=32.51 (0.027)
C. Silva et al. / Environment International 47 (2012) 86–98
farms may lead to a stress on benthic fauna population dynamic in-
cluding a loss of biodiversity, number of species and abundances.
Although biomarker approach did not show a clear pattern, the bio-
chemical responses in clams that reflect oxidative stress (GPX and
LPO) and DNA damage, better reflected sediment contamination
and can be indicative of the actual environmental risk and health sta-
tus of the exposed organisms.
The effects of organic loads from fish aquaculture activities in the
sediments and benthic communities of some European coastal areas
are well documented (Kalantzi and Karakassis, 2006; Karakassis et
al., 2002; Mantzavrakos et al., 2007; Papageorgiou et al., 2009;
Pusceddu et al., 2007; Tomassetti and Porrello, 2005; Vezzulli et al.,
2008). The accumulation of organic matter in the substrate depends
on the oxidative potential of sediment for nutrient remineralisation
(Nixon, 1981). If organic matter sinks are mineralised by dissolved
oxygen in the sediment–water layer, organic enrichment stimulates
benthic fauna biodiversity by providing trophic resources (Nickel et
al., 2003). However, if organic matter sinks are not mineralised by
dissolved oxygen, the results are hypoxia or even anoxia, with a dete-
rioration of the health status of benthic fauna and a loss of biodiversi-
ty (Pearson and Black, 2001). Ammonia, sulphides and methane are
the main toxic compounds derived from the anaerobic mineralisation
of organic matter (Thamdrup and Canfield, 2000).
In this study a benthic organic enrichment in the surroundings of a
fish farm effluent following a pollution gradient was evidenced and
accompanied by a change in the sediment characteristics. A spatial
pollution gradient was observed in TOM with higher values in the
upper part of the creek in the sampling sites close to the fish farm
effluent. Significant (pb0.01) negative correlations were estimated
between TOM, pH and Eh. Lower values of Eh and pH were observed
in the surrounding areas of the fish farm with significant (pb0.01) dif-
ferences from the control site. Eh values indicate that a trend toward
reduction conditions would be due to the effects of the fish farm efflu-
ent on the seabed. Eh is a qualitative metric of the intensity of reduc-
tion conditions (SEPA, 2005) that could reflect organic enrichment,
but low Eh could also be the result of natural physical processes, espe-
cially in soft sediments (Brooks et al., 2002). Therefore, its usefulness
in assessing the effects of fish farms is limited if it is not accompanied
by a great knowledge of the study area. To correctly measure Eh in
sediments the sampling method must ensure that the vertical profile
is not altered; this was guaranteed by the in situ sampling in the in-
tertidal muddy sediment during neap tide. Organic enrichment of
the seabed involves sediment anoxia, increased nutrients, deteriora-
tion of the optical properties of seawater and impoverished sediment
quality due to organic matter sedimentation (Holmer et al., 2005).
The fact that high (>10%) TOM content (Long and Morgan, 1990) is
associated with sediment anoxia is supported by the significant
(pb0.01) negative correlation with DO concentration estimated in
The spatial trend and pollution gradient in the sediment charac-
teristics obtained in this study is consistent with previous results on
the longitudinal distribution of various water quality parameters in
the RSP creek (Tovar et al., 2000a, b). The higher levels of TOM and
lower values of pH, Eh and DO observed in this research in sediments
closed to the aquaculture effluent are consistent with the increased
values of nutrients, particulate organic matter and suspended solid
and decreased pH observed in water samplings by Tovar et al.
(2000a, b). Additionally, water temperatures observed in October
2010 (autumn) were similar to those obtained by Tovar et al.
(2000a) in October 1996, with values varying between 20 and 23°C.
However, salinity values were lower than those obtained by Tovar
et al. (2000a), may be due to the higher precipitation (89mm)
recorded during October 2010 compared to the value (55mm) regis-
tered in October 1996 by the Spanish Meteorological Agency in Jerez
de la Frontera airport (36°44′45″S, 06°03′48″W) station located
28km from the study site (AEMET, 2012).
Effects of stress on population dynamics include: increase in produc-
ber of species with increasing number of individuals (Pearson and Black,
2001). Benthic faunal studies have been used to assess in situ alterations
in residential community structure in relation to pollution-induced
changes derived from aquaculture activities (Pusceddu et al., 2007;
Tomassetti and Porrello, 2005; Vezzulli et al., 2008). In RSP, this benthic
community analysis has not been difficult due to the existence of an ap-
in an area that does not receive any recognised pollution inputs and has
good water circulation. It is clear that the benthic community has been
affected by the contamination of the fish farm effluent located in the
upper part of the creek. Sediment alteration in the more contaminated
sites (RSP1, RSP2 and RSP3) was due to an increase in fish farm waste
and low water renovation, allowing accumulation of organic matter,
oxidative conditions and anoxia in the area of the sea floor. In situ alter-
a spatial gradient of contamination from the aquaculture effluent to the
control site. Benthic community at the more contaminated sites with re-
spect to the control site, show perturbations caused by organic enrich-
ment: fewer species in sampling sites near the effluent, lower number
of individuals and diversity indexes, and MDS and CCA plots showing
(RSP1, RSP2 and RSP3) sites. The analysis performed across spatial scale
provides additional insights into how benthic fauna respond at the
different sites following the pollution gradient. The PERMANOVA analy-
sis revealed highly significant differences in benthic fauna structure
among the control site and the site close to the aquaculture effluent,
which suggests that the effects of fish farming vary across the pollution
In the present study the most common and abundant species at
the clean or control site, which dominate the intertidal benthic com-
munity in muddy sediments, were similar to those reported in previ-
ous studies for similar unpolluted ecosystems in the Gulf of Cádiz and
Portugal, including mollusks (S. plana and H. ulvae), oligochaetes and
polychaetes (P. lyra, N. diversicolor) (DelValls et al., 1998; França et al.,
2009; Rodrigues et al., 2006). The predominant benthic fauna found
at the most contaminated sites were isopods C. carinata, polychaetes
and nematodes. These species pointed out the presence of a similar
Condition factor (g mm-1)
Fig. 4. Condition factor (total weight/shell length) of Scrobicularia plana collected in
five sampling sites at the Río San Pedro creek, SW Spain. Error bars indicate standard
deviations based on samples of 30 clams at each site. Asterisks (*) indicate significant
induction (pb0.05) compared with the control site RSP5.
C. Silva et al. / Environment International 47 (2012) 86–98
community to the one described as ‘Reduced Macoma’ by Thorson
(1957) and characteristic of estuary sites with salinity stress. How-
ever, oligochaetes, which are a group of taxa considered more tolerant
to pollution, were not found in the more contaminated sites (DelValls
et al., 1998). Additionally, the soft bottom benthic community affected
by fish farm effluents shows a prevalence of small species (Heilskov
et al., 2006; Macleod et al., 2004) in line with the Pearson and
Rosemberg (1978) model. This model predicts a reduced mean body
size, a shallower distribution and an impoverished functional commu-
nity structure with increasing organic load (Heilskov et al., 2006).
CCA and BIOENV analyses indicated that a relatively large propor-
tion of among-site variances in abundance of benthic fauna along the
pollution gradient was strongly associated with the environmental
variables measured, particularly, sediment-associated redox poten-
tial, pH, dissolved oxygen and organic matter. Changes in benthic
fauna abundance and diversity were related to a decrease of redox
potential and pH in the muddy sediment's stations close to fish
farms, indicating a possible hypoxia/anoxia situation (Tomassetti et
al., 2009). Hypoxia and anoxia causes mortality in many invertebrates
and is one of the major factor altering benthic fauna in areas impacted
by fish aquaculture (Gray et al., 2002; Tomassetti et al., 2009). CCA
analysis also demonstrated similar patterns in the taxa associated to
the most contaminated sites; in high organically enriched sediments,
the faunal community is composed of small fauna (isopods, poly-
chaetes and nematodes) tolerant of these environmental conditions.
mol min- -1mg-1
RSP1RSP2 RSP3 RSP4RSP5
RSP1RSP2RSP3 RSP4 RSP5
RSP1 RSP2RSP3 RSP4RSP5
RSP1RSP2 RSP3RSP4 RSP5
µmol mg -1
Fig. 5. Oxidative stress parameters, antioxidant and detoxification enzymatic activities in the digestive glands of captured in five sampling sites at the Río San Pedro creek, SW Spain.
Ethoxyresorufin O-deethylase (EROD), glutathione S-transferase (GST), glutathione peroxidase (GPX) and glutathione reductase (GR) activities; Lipid Peroxidation (LPO) and DNA
damage. Error bars indicate standard deviations of mean based on 5 replicates at each site. Asterisks (*) indicate significant induction (pb0.05) compared with the control site RSP5.
Axis 2 (23.53 %)
Axis 1 (67.35 %)
(Axis 1 and 2: 90.88 %)
Fig. 6. CCA triplotshowingscoresof sites,the 15mostabundantspeciesandenvironmen-
tal variables. Full species names are given in Table 3. T = temperature; S = salinity; Eh =
redox potential; DO = dissolved oxygen; and TOM = total organic matter.
BIOENV correlations (Spearman ran correlation coefficients) between ben-
thic community structure and environmental variables. Sample statistic
(Rho): 0.882, pb0.001, Number of permutations: 999 (Random sample).
Single variable Variable combination
Eh, pH, TOM (0.869)
pH, TOM (0.869)
SA, Eh, pH, TOM (0.826)
SA, T, Eh, pH, TOM (0.825)
T, Eh, pH, DO, TOM (0.796)
SA, Eh, pH, DO, TOM (0.74)
C. Silva et al. / Environment International 47 (2012) 86–98
At the molecular or cellular level, the strategies of cellular protec-
tion and repair in response to stress and harmful conditions (e.g. high
or low temperature, ultraviolet light, oxidative conditions, anoxia and
salinity stress) comprise oxidative damage, alterations in biotransfor-
mation and antioxidant enzymes (Stegeman et al., 1992; Winston and
Di Giulio, 1991). In the present study stress biomarkers were mea-
sured in the DG of clams because it is the most sensitive tissue to
sediment pollution (Banni et al., 2009; Barreira et al., 2007; Zhang
et al., 2010). Study sites are affected mainly by the contaminants dis-
charged by the fish farm effluent. However, elucidating the origin of
an observed biological alteration in an environment is difficult due
to the range of possible causes. The biomarkers selected in this
study are recommended for pollution monitoring studies (Gagné et
physico-chemical measurements and EROD, GST and GR enzymatic ac-
tivities were found in this research. In situ biological alterations were
reflected by significant changes in GPX, LPO and DNA strand break
biomarker responses in S. plana, following a spatial gradient of contam-
ination from the fish farm effluent to the control site. In the more con-
taminated sites, the aforementioned biomarkers showed significant
enhancements in comparison with the control site. In this study,
S. plana was used as an adequate sentinel bivalve for pollution bio-
monitoring and this specie was present at all the sampled sites due to
the muddy characteristics of the sediments. There are only a few recent
studies using biomarkers in S. plana (Coelho et al., 2008; Romero-Ruiz
et al., 2008; Solé et al., 2009).
EROD activity was induced in clams collected in sediments from
the most contaminated sites of the RSP. However, this induction
was not significant. No significant induction of EROD activity can be
explained by the acute hypoxia and because the RSP is located in a
dead mouth river that does not receive any recognised organic con-
tamination. EROD activity is involved in the first phase of metabolism,
unmasking or adding reactive functional groups during hydrolysis
oxidation and reduction (Goeptar et al., 1995). Induction is a clear
sign of CYP1A1 enzyme activity, which is among the many known cy-
tochrome isoforms mainly involved in Phase I organic xenobiotic bio-
transformations. EROD activity has been reported in many species of
invertebrates, including clam species (Martín-Díaz et al., 2008;
Pereira et al., 2010; Pérez et al., 2004; Solé and Livingstone, 2005),
following exposure to organic trace pollution (Lafontaine et al.,
2000). The involvement of xenobiotics (PCBs, PAHs, pharmaceuticals)
in EROD induction was reported by different authors (Devier et al.,
2005; Martín-Díaz et al., 2009; Ramos-Gómez et al., 2011; Van der
Oost et al., 2003). Additionally, acute hypoxia has been observed to re-
duce the rate of biotransformation by CYP1A1 (Fradette et al., 2007).
GST activity did not significantly increase in clams collected in sedi-
ments from the most contaminatedsites of the RSP. GST is a Phase II bio-
transformation enzyme whose activity has been reported to be induced
in bivalves under pharmaceutical (Martín-Díaz et al., 2009) and PAH
(Regoli et al., 2002) induction. A previous study has reported inhibitory
effects of metals on GST activity in S. plana from the Bay of Cadiz (Solé
et al., 2009). Metals have been responsible for enhanced reactive oxygen
enzymes and GSH content in bivalves (Canesi et al., 1999; Regoli et al.,
1998; Solé et al., 2009). Moreover, higher levels of TOM and food avail-
ability for clams found at the most contaminated sites can account for
changes in GST activities (Bocchetti and Regoli, 2006; Sáenz et al.,
2010). A decrease in GR can cause a decrease in GSH, which leads to de-
ity, denoting the importance of the coordinated regulation of those
enzymes in maintaining the equilibrium between the cellular GSH and
in the maintenance of GSH levels, which is used in turn as a substrate by
GST (Reed, 1986; Sies, 1991). However, the increases in GR activity in
clams collected in sediments from RSP3 site could be due to exposure
to other chemicals that are present in this sampling area.
GPX activity was significantly increased in clams collected in
sediments from the most contaminated sites, indicating a defence
mechanism against oxidative stress. GPX is involved in the inhibition
of ROS oxyradical formation, reducing H2O2to H2O with GSH oxida-
tion. Induction of GPX provides an indication of higher H2O2produc-
tion in the DGs of clams from the sites close to the fish farm,
suggesting the presence of redox-active contaminants. Increased ac-
tivities of GPX in bivalves have been described in previous studies,
particularly in clams, oysters, scallops and mussels (Faria et al.,
2010; Martín-Díaz et al., 2007, 2008; Niyogi et al., 2001; Pereira et
al., 2010; Pérez et al., 2004; Regoli et al., 1998).
Sediments from the more contaminated sites were characterised by
lower pH and Eh and higher organic matter, oxidative conditions and
hypoxia. LPO as well as DNA strand breaks are well-known effects of
oxidative stress (Valavanidis et al., 2006) provoked by the action of
contaminants and/or their metabolites through molecule binding or
ROS production. ROS can be highly toxic to aquatic organisms, often
resulting in the oxidation of lipid membranes, protein oxidation and
DNA damage (Almeida et al., 2007). Organisms have developed a de-
fence system that includes enzymatic antioxidants and non-enzymatic
antioxidants, whose function is to keep ROS at metabolically innocuous
breaks (VanderOostetal.,2003).LPO andDNA strandbreaksexhibited
the pollution gradient. These results are in accordance with other
studies reporting the role of LPO as a biomarker of oxidative damage
in cells and tissues of bivalves (Pampanin et al., 2005; Ramos-Gómez
et al., 2011). Oxidative stress effects in aquatic organisms have been
described in previous studies associated with hypoxia and acidic sedi-
ment conditions (Vidal et al., 2002) and with the use of ozone (Ritola
et al., 2000, 2002), hydrogen peroxide (Rach et al., 1998) and other
pro-oxidant chemicals in aquaculture (Livingstone, 2003). However,
there is a lack of studies concerning the oxidative effects related to the
pollution of the discharged effluents of aquaculture farms.
An important aspect related to the use of a biomarker approach in
biomonitoring of coastal areas is the need of having a detailed knowl-
edge of natural variations of confounding factors contributing to their
response and distinguish pollution induced effects from those induced
by natural biological cycle of clams, including the reproductive cycle
and spawning period (Bocchetti et al., 2008; Hagger et al., 2006). Mor-
phological CF of S. plana varies between sites, although a narrow size
range of analysed individuals was used intentionally selected in order
to avoid size effects on biomarker responses. The CF was significantly
increased at sites (RSP1 and RSP2) nearest to the fish farm effluent
and this spatial pattern may be due to the gain of weight/body mass
associated with higher food availability for clams, specifically chloro-
phyll a and particulate organic matter in water which influence the as-
similation efficiency of shellfish filters (Gangnery et al., 2003). In fact,
this study obtained higher values of TOM in sediments close to the
aquaculture farm compared to control site, which are consistent with
the spatial gradient with higher values of chlorophyll a and particulate
organic matter nearest to farm effluents observed by Tovar et al.
(2000b). However, no significant correlations were found between
the environmental variables, biomarker responses and the clam's CF,
suggesting that this morphological cofounding factor was marginally
related to the biochemical responses of biomarkers. The reproductive
cycle of S. plana from the RSP creek, should be similar of those
from the estuary of Guadalquivir (Cádiz, SW Spain) river with a clear
spawning period from March to September and a period of sexual re-
pose from October to January (Rodríguez-Rúa et al., 2003). The reduced
GSI values obtained during October in the RSP in the present study
were consistent with the reproductively inactive phase of this clam
for similar coastal areas in Spain (Rodríguez-Rúa et al., 2003; Sola,
1997). However, the GSI measured in clams of the RSP indicated signif-
icant spatial differences with higher values in the control site (RSP5),
suggesting that higher energy-rich gonad (more nutritive value) and
C. Silva et al. / Environment International 47 (2012) 86–98
advanced reproductive status are associated with the control site with
low aquaculture impact. The decrease in GSI indicates an overall degra-
dation of animal health and may be attributed to the higher concen-
trations of TOM and lower levels of Eh and DO in sediment of the
contaminated sites compared to control site. In fact, decreased repro-
ductive capability in feral organism may be considered as one of the
most damaging effects of persistent anthropogenic pollutants (Van
der Oost et al., 2003). The GSI and other several effects of the so-called
endocrine-disrupting xenobiotics on organism reproduction have
been chosen as biomarker (Gagné and Blaise, 2005). However, no sig-
nificant correlations were found between the biomarker responses
and the clam's GSI in this study. Several works have demonstrated
thatoxidativestressparameters(LPOandDNAstrand breaks), detoxifi-
cation enzymes (EROD and GST) and antioxidant enzymes (GPX and
GR) of bivalves present higher values during spring and summer
being associated with the spawning stage (Bocchetti and Regoli, 2006;
Pereira et al., 2010; Ramos-Gómez et al., 2011; Verlecar et al., 2008).
Sincethis study wasconducted in autumn when clams were at the sex-
ual repose stage with a low GSI, a lower biomarker induced responses
maybeattributed to thereproductionstagefactor. Nevertheless, higher
GSI values in the control site (RSP5) compare to other sites may induce
higher biomarker responses and this could generate overestimation
values. EROD induction is known to vary with season and reproductive
status (Pereira et al., 2010) and this could explain the great values in
RSP5 associated with higher GSI values. Also a possible overestimation
of the oxidative stress parameters and antioxidant biomarkers in RSP5
possibly associated to higher GSI should be considered when assessing
the reasons for observed variability. Additional confounding factors to
those presented in this study, such as biological (sex), environmental
(other physico-chemicals) and seasonal variability, must be carefully
taken into account in future biomarker assessments. The utility of
some biomarkers in detecting interregional difference may be con-
strained because of natural variations of confounding factors contribut-
ing to their response (Hagger et al., 2006).
Anoxia, reduced potential, acidic and high organic enrichment
conditions were registered in sediments at the most contaminated
sites, suggesting a strong association with the observed lower levels
of benthic biodiversity indicators (S, N, d, H′). It has been proven
that organic enrichment of sediments impacted by fish farms can
be the cause of the decreased biodiversity observed, which follows a
spatial gradient of contamination. Sediment conditions observed in
the most contaminated sites were also correlated with the observed
oxidative stress responses in clams: higher GPX enzymatic activity,
LPO and DNA damage. Metabolic activities, and consequently the ox-
idative stress status of bivalve's species, are subject to spatial varia-
tions due to biotic and abiotic factors. The results suggest that
organic enrichment of sediments impacted by fish farms can be an
important source of pro-oxidants in clams, leading to the production
of radicals (Lima et al., 2006). It was demonstrated that acute hypoxia
and acidification in sediments increase the production of ROS, leading
to oxidation of cellular components.
Rio San Pedro presented a pollution gradient from the fish farm
effluent to the control site characterised by hypoxia/anoxia, reduced
potential, acidic conditions and high organic enrichment in sediments
at the most contaminated sites.
The present study revealed a decrease in biodiversity indicators
(S, N, d, H′) of the benthic community of intertidal sediments along
the pollution gradient. This decrease was produced by the effects of
discharged fish farm effluents, which is a strong indication that severe
chronic effects exist in this shallow tidal creek.
Biomarker responses in S. plana did not show a clear pattern and
few correlations with sediment conditions, however significant dif-
ferences between the most contaminated sites close to aquaculture
discharges versus the control site were obtained, which could be
attributed to the induction of oxidative stress (LPO and GPX activity)
and DNA damage in clams and may lead to an alteration of the health
status of the exposed organisms. It has been proven that organic en-
richment of sediment impacted by fish farm effluents are sources of
pro-oxidants in clams, leading to ROS production, oxidation and dam-
age of cellular components. S. plana seems to be an adequate sentinel
in the ecosystem.
sediment conditions along the pollution gradient, causing biodiversity
and biochemical changes and thus potential dangers for benthic organ-
isms. It has been demonstrated that effluents from fish aquaculture ac-
lead to analteration of thehealthstatusand biodiversityof theexposed
Biomarker responses can be used as early warning tools of the po-
tential adverse effect in organisms, while the benthic fauna population
with ecological and ecosystem relevance.
C. Silva is grateful to the Sistema Bicentenario BECAS CHILE
from the Chilean Government and the international grant from Bank
Santander/UNESCO Chair UNITWIN/WiCop for funding this work.
The authors are grateful to the anonymous reviewers for their valu-
able comments and suggestions to improve the quality of the paper.
This is the CEIMAR journal publication 6.
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