Post on 30-Jan-2023
Chemosphere 52 (2003) 1209–1224
www.elsevier.com/locate/chemosphere
Integrative assessment of marine pollution in Galicianestuaries using sediment chemistry, mussel
bioaccumulation, and embryo-larval toxicity bioassays
R. Beiras a,*, N. Fern�aandez a, J. Bellas a, V. Besada b,A. Gonz�aalez-Quijano b, T. Nunes b
a Departamento de Ecolox�ııa e Biolox�ııa Animal, Universidade de Vigo, E-36200 Vigo, Galicia, Spainb Instituto Espa~nnol de Oceanograf�ııa (IEO), Cabo Estay, Canido, E-36200 Vigo, Galicia, Spain
Received 13 March 2002; received in revised form 10 January 2003; accepted 28 March 2003
Abstract
An integrative assessment of environmental quality was carried out in selected sites along the Galician coast (NW
Iberian Peninsula) combining analytical chemistry of seawater and sediments, bioaccumulation in the marine mussel,
and embryo-larval sediment toxicity bioassays, in order to link biological and chemical criteria for the assessment of
coastal pollution. Maximum values of Hg and Cu in seawater, sediment and mussels, were found in the inner part of
Ria of Pontevedra, while maximum levels of organics (polychlorinated biphenyls, hexachlorobenzene and aldrin) were
found in mussels from A Coru~nna. Outstanding values of Cu, Pb and Zn have been found in seawater and sediment from
a single site, P3, which also was the most toxic in the embryo-larval bioassays performed with four different phyla of
marine organisms: mollusks, echinoderms, arthropods and chordates. Sediment quality effects range-median values
provided a valuable reference to predict biological effects from sediment chemistry data, while effects range-low values
were too conservative. Sediment toxicity could also be predicted by using a toxic-unit model based on published EC50
values for trace metals and mobilization factors independently obtained from measurements of metal contents in
sediments and their elutriates. When chemical and toxicological data are independently used to arrange sampling sites
by using non-metric multidimensional scaling, a remarkable degree of concordance between both types of configura-
tions could be observed.
� 2003 Elsevier Science Ltd. All rights reserved.
Keywords: Paracentrotus lividus; Mytilus galloprovincialis; Ciona intestinalis; Trace metals; Chlorinated organics; Sediment toxicity
bioassays
1. Introduction
Marine sediments and mussels are commonly used as
environmental matrices in chemical monitoring pro-
grams because they accumulate persistent pollutants at
concentrations orders of magnitude above those in the
*Corresponding author. Tel.: +34-986-812648; fax: +34-986-
812556.
E-mail address: rbeiras@uvigo.es (R. Beiras).
0045-6535/03/$ - see front matter � 2003 Elsevier Science Ltd. All ri
doi:10.1016/S0045-6535(03)00364-3
water (e.g. OSPAR Commission, 2000). Mussels have
been selected as indicator organisms world wide because
of their abundance, ubiquity, sessile nature, long life
span, high filtration rates, and especially, because of
their high bioconcentration factor for most pollutants of
concern in the sea (Whitfield, 2001). Biological assays
are currently used as a rapid and cost-effective screening
tool to detect polluted sites where thorough analytical
chemistry may be subsequently performed. In addi-
tion, bioassays respond to the bioavailable fraction
of the toxicants, thus providing ecologically relevant
ghts reserved.
Fig. 1. Map of the sampling sites in the Galician Rias, and
location of the Rias in the Atlantic coast of Europe (insert).
1210 R. Beiras et al. / Chemosphere 52 (2003) 1209–1224
information. Because of their high sensitivity, early life
stages of marine invertebrates are particularly suitable
for marine environmental bioassays. First reported by
Wilson (1951), sea-urchin embryo development may be
impaired by waterborne toxicants, indicating a poor
biological quality of that water sample. This concept is
the basis of later assessment of coastal environmental
quality using sea-urchin (e.g. Kobayashi, 1971, 1991;
Bougis et al., 1979; Vashchenko and Zhadan, 1993; Carr
et al., 1996; Beiras et al., 2001) and bivalve (e.g. Woelke,
1961; Cardwell et al., 1977; Thain, 1992; His and Beiras,
1995) embryogenesis and larval development (for a re-
view see His et al., 2000). Recently, tunicates were also
proposed in order to detect toxicants which may act
selectively on chordates (Bellas et al., 2001).
In order to extract meaningful information from the
large and heterogeneous bulk of data generated by the
chemical analyses and toxicological assays, multivariate
statistical methods are currently employed (e.g. Shin and
Fong, 1999; Nascimento et al., 2001). Among the mul-
tivariate techniques, non-parametric tests are particu-
larly suitable to deal with non-linear responses, such as
those recorded in toxicological assays, and with highly
variable variables that contain only noise, which may be
the case for the levels of certain chemicals. Non-metric
methods do not rely on data transformation or unwar-
ranted assumptions on their theoretical distribution, and
their use have been recommended within the field of
ecotoxicology (Landis et al., 1997).
The goal of the present study was to assess coastal
marine pollution in the main Galician estuaries by
combining chemical and toxicological data in order to
compare and integrate both approaches. Toxicity bio-
assays provide ecologically meaningful information on
the threat posed by pollution, while analytical chemistry
contributes to the interpretation and explanation of
toxicity patterns. The information acquired will allow
ordination of the sampling stations by using non-metric
multidimensional scaling (MDS) and classification into
clusters based on similarity of both chemical and toxi-
cological data.
2. Materials and methods
2.1. Field sampling, environmental parameters and ana-
lytical chemistry
2.1.1. Sampling sites
On spring 1997 and 1998 we sampled three sites in
each of the three major Galician Rias (NW Iberian
Peninsula): Ria of Vigo (V 1–3), Ria of Pontevedra (P 1–
3), and Ria of Arousa (A 1–3). The sites (Fig. 1) were
chosen at the inner part of the Southern coast of each
estuary, representing some of the most polluted places in
the Rias according to previous data, and the time rep-
resented the seasonal maxima of pollutants in mussels
for this area (Fumega et al., 1984). On 1998 we sampled
in addition two sites at a fourth estuary, Ria of O Burgo
(B 1–2), in the vicinity of a major Galician city (A
Coru~nna). Sampling dates and locations are shown in
Table 1.
2.1.2. Water samples
Triplicate samples of 250 ml were taken in poly-
propylene flasks and stored at )20 �C for analysis of
phosphate. These flasks were previously washed with
diluted HCl and bleach, and rinsed with distilled water.
Triplicate pairs of samples of 140 ml were taken in dark
bottles for the measurement of biological oxygen de-
mand (BOD5) from the oxygen consumption after 5 d
incubation at 20 �C. These bottles were previously wa-
shed with diluted nitric acid. Dissolved oxygen was
measured by the Winkler method (APHA-AWWA-
WPCF, 1992) The detection of the titration endpoint
was visual in 1997 and potentiometric, by using a 721
NET Titrino (Metrohm), in 1998. Triplicate samples of
1 l were taken in polypropylene flasks, acidified in situ
with 1 ml/l of ultrapure nitric acid, and stored at room
temperature for analysis of Cd, Cu, Pb and Zn. These
Table 1
Locations and dates of sampling
Estuary Site Location (N–W) Sampling dates
Ria of Vigo V1 42�1702100–8�3702800
08/04/97 26/03/98
V2 42�1701200–8�3705900
V3 42�1702500–8�3802500
Ria of Pontevedra P1 42�2505700–8�3805400
08/05/97 27/03/98
P2 42�2402900–8�4005700
P3 42�2402500–8�4101700
Ria of Arousa A1 42�3802500–8�4503700
22/05/97 31/03/98
A2 42�3703000–8�4601000
A3 42�370000–8�4602900
Ria of O Burgo B1 43�1805900–8�2103000
– 30/03/98
B2 43�1904800–8�2202000
B sites were sampled in 1998 only.
R. Beiras et al. / Chemosphere 52 (2003) 1209–1224 1211
flasks were previously washed with ultrapure acids.
Surface water temperature, salinity, pH and dissolved
oxygen were recorded in situ by means of Orion and
Hanna electrodes.
2.1.3. Sediment samples
The surface layer (2 cm) of the intertidal sediments
was collected with wooden palettes at low tide, placed
into sealed polyethylene bags, carried to the laboratory
on ice and stored at 4 �C in the dark. Samples intended
for chemical analyses were stored freeze-dried in glass
flasks. Redox potential at 1, 3 and 5 cm depth was re-
corded in situ by means of a Crison electrode. Organic
matter content was measured from triplicate sediment
samples by drying at 80 �C until constant weight, and
incineration in furnace at 450 �C for 24 h. Salt weight
was discounted from the sediment dry and ash weight.
Mean particle diameter was calculated from the dry
weight of the fractions after sieving at 4, 2, 1, 0.5, 0.25,
0.125 and 0.063 mm (Restch, AS2000 model).
2.1.4. Mussel samples
Mussels (Mytilus galloprovincialis) of approximately
4 cm long were sampled from each site (with a single
exception, P1, where the fresh water prevented the oc-
currence of this species), placed into sealed polyethylene
bags and stored at )20 �C.
2.1.5. Analysis of metals in seawater
Heavy metal solutions were prepared from 1 g/l (in
0.5 N HNO3) Atomic Absorption Standards. For Cu,
Zn, Cd and Pb voltammetric measurements were carried
out by means of a GSTAT10 (Eco-Chemie B.V.) po-
tentiostat coupled to a VA Stand 663 (Metrohm). A
hanging mercury drop electrode of 0.33 mm2 of area was
used as working electrode, and potentials were measured
versus a saturated calomel reference electrode. The
voltammetric procedure consisted of anodic stripping
voltammetry (ASV). A differential pulse scanning mode
was chosen with a scan speed of 10 mV s�1 and a su-
perimposed pulse amplitude of 50 mV. Purge time,
stirring, deposition time, and other typical variables
were controlled with a General Purpose Electrochemical
System version 4.2 software (Eco Chemie B.V.). Nitro-
gen (N50) was used to deaerate solutions. The electro-
chemical cell was constructed of borosilicate and had
been previously siliconised. Before each analysis, the cell
was thoroughly washed with a sulfonitric solution con-
taining potassium dichromate to saturation. Afterwards
it was fully flushed with plenty of Milli-Q50 (Millipore,
Waters) purified water. Natural water samples were fil-
tered through a 0.45 lm cellulose acetate filter. Hydro-
gen peroxide 30% (Analysis grade), and nitric acid 60%
(Analysis grade) were used for digestion of organic
matter. An in-line UV photochemical digestion of
samples was achieved by means of a low-pressure 14 W
mercury lamp that axially irradiated a 15 ml quartz coil,
through which the sample flowed, forced by a peristaltic
pump. Manipulation of samples, filtration procedures,
reagent additions and cell solution preparation were
carried under laminar flow of air in a hood (Crumair)
provided with a pre-filter and an ULPA 0.2 mm absolute
final filter.
2.1.6. Analysis of metals in sediment and mussels
The digestion of total sediment (fraction <2 mm;
ICES, 2000) were digested with a mixture of HNO3–
HCl–HF in teflon digestion bombs by means of a mic-
rowave oven. The excess of HF was neutralized with
HBO3. The Cu, Zn, Ni, Fe and Li were determined by
atomic absorption spectrometry (Perkin–Elmer Zeeman
5000 with HGA 400) with air–acetylene flame and Pb,
Cd, As and Cr in graphite furnace with Zeeman back-
ground correction (Besada and Schultze, 1999). Mussel
samples were digested with HNO3 in teflon bombs and
analyzed by atomic absorption spectrometry with the
same equipment as for sediments. Mercury was also
analyzed in the seawater, sediment and mussels; the
1212 R. Beiras et al. / Chemosphere 52 (2003) 1209–1224
methods and results were reported elsewhere (Beiras
et al., 2002), and those data will be used here only for the
correlations with other pollutants and multivariate
analysis.
2.1.7. Analysis of organics in sediments and mussels
Analytical methods for individual polychlorinated
biphenyls (PCBs), hexachlorobenzene (HCB), aldrin
and 4,40DDE followed for biota De Boer (1988) and
Gonz�aalez-Quijano and Fumega (1996), and for sedi-
ments (fraction <2 mm) Smedes and De Boer (1997),
and Nunes et al. (1997). Samples were extracted in a
Soxhlet apparatus with a mixture pentane–dichloro-
methane (1:1 vol.). The extracts were purified in an alu-
mina column eluted with pentane, and fractionated in a
silica gel column sequentially eluted with iso-octane
and iso-octane/diethylether. In addition, sulfur was
previously eliminated from sediment samples by adding
an aqueous solution of sodium sulfite and tetrabutyl-
ammonium (Jensen et al., 1977; Nylund et al., 1992).
The purified extracts were analyzed by gas chromato-
graphy with electron capture detector (Perkin–Elmer
PE 8500 and Autosystem) using capillary columns (50
m, 0.22 mm i.d. and 0.33 lm film) and H2 as carrier
gas. The following seven individual PCBs were ana-
lyzed: IUPAC numbers 28, 52, 101, 118, 138, 153 and
180, considered as priority marine pollutants by inter-
national agencies.
2.1.8. Quality control
The quality of the chemical analyses at IEO-Vigo was
controlled at two levels: (i) Internally by using certified
reference materials along with each series of samples.
For sediments, HS-1, BCSS-1 and BEST-1 reference
materials were obtained from the National Research
Council of Canada. For mussels, the reference materials
CRM-278, CRM-349 and CRM-350 were obtained
from the BCR (Boureau Communautaire de Reference,
EU). (ii) Externally, by participation in intercalibration
exercises promoted by international institutions, namely
the International Atomic Energy Agency (IAEA), (Co-
query et al., 2000; Villeneuve and de Mora, 2000), the
QUASIMEME programme (Quality Assurance of In-
formation for Marine Environmental Monitoring in
Europe) from 1993 to 1996 (Wells, 1996), QUASI-
MEME Laboratory Performance Studies from 1996 to
date (QUASIMEME, 2001), and QUASH (Quality
Assurance of Sample Handling) from 1997 to 2000
(Smedes et al., 2000). IEO-Vigo was Reference Labo-
ratory for the metal analyses in sediments within
QUASH.
2.2. Embryo-larval bioassays
The toxicity of the sediments from all sampling sites
was assessed in 1997 and 1998 by means of sediment
elutriate bioassays with embryos of the echinoderm
Paracentrotus lividus. Additional bioassays with selected
samples were performed using early developmental
stages of the bivalve M. galloprovincialis, the crustacean
Palaemon serratus, and the chordate Ciona intestinalis.
These species are hereafter referred to as sea-urchin,
mussel, prawn, and ascidian, respectively.
Elutriates were obtained according to methods pre-
viously described (Beiras and His, 1995; Beiras, 2001),
consisting of magnetic stirring (100 g sediment in 400 ml
seawater) for 30 min and decantation for 12 h approxi-
mately, at 20 �C. Long decantation times allowed testing
unfiltered elutriates. Elutriates from each sediment
sample were tested undiluted (1:4 sediment:water w:w),
and after dilution to 1:16 and 1:40 sediment:water (w:w).
Water for dilutions and blanks was 0.22 lm-filtered
oceanic seawater in 1997 and artificial seawater (His
et al., 1997) in 1998. Temperature, salinity, pH and
dissolved oxygen in the elutriates was recorded prior to
the bioassays. When the presence of hydrogen sulfide in
the elutriates was suspected, its concentration was also
measured by means of a commercial colorimetric test
(Merk, Aquaquant), after 5 times dilution of elutriate
samples in distilled water.
Sea-urchin bioassays were based on the methods
described by Fern�aandez and Beiras (2001). Gametes
were obtained by dissecting mature individuals collected
from natural populations inhabiting the outer part of
the Ria of Vigo. Gamete quality (round eggs and motile
sperm) was checked under the microscope and one fe-
male and one male showing optimal condition were se-
lected. Eggs from the selected female were transferred
into a 100 ml measuring cylinder containing 0.2 lm fil-
tered seawater, a few ll of undiluted sperm from the
selected male were added and the contents of the cylin-
der were gently stirred with a plunger for 2 min ap-
proximately to allow fertilization. Four 50-ll samples
were then taken by pipette and observed under the mi-
croscope in order to record the number of eggs and
percentage of fertilization. Fertilization success (indi-
cated by the presence of a fertilization membrane) was
generally higher than 98% in all the batches used for the
bioassays. Within 30 min after fertilization volumes
corresponding to 600 fertilized eggs were transferred
into 20 ml vials containing the elutriates. Five replicates
of each experimental treatment, including the blank,
were used (except Pontevedra 1997, where n ¼ 3). The
vials were incubated for 48 h at 20 �C, fixed by adding
two drops of 40% formalin, and directly observed under
an inverted microscope. Embryogenesis success, evalu-
ated as percentage of 4-arm pluteus larvae, and larval
length were recorded.
Mussel bioassays followed His et al. (1997), and
consisted of incubating 30 min-old fertilized eggs (25 per
ml) in the experimental solutions at 20 �C for 48 h. The
biological response recorded was percentage of D-larvae
Fig. 2. Embryogenesis success for sea-urchin exposed to elutriates from sediments sampled in Vigo (a), Pontevedra (b), Arousa (c) and
O Burgo (d) in 1998. Symbols indicate sites 1 (diamonds), 2 (squares) and 3 (triangles).
R. Beiras et al. / Chemosphere 52 (2003) 1209–1224 1213
with normal shell. This excludes shell abnormalities ((b),
(c) and (d) in Fig. 2, His et al., 1997) but not protruding
mantle ((e) in Fig. 1 op. cit.). Ascidian bioassays fol-
lowed Bellas et al. (2001); 1 h-old fertilized embryos (2
cell stage) were incubated at 20 �C and a density of 15
per ml for 24 h, and percentage of normal tadpole larvae
was recorded. Prawn bioassays followed Mari~nno-Balsa
et al. (2000); newly hatched larvae (zoea I) were incu-
bated at 18 �C and a density of 10 individuals in 25 ml
for 72 h, and mortality (assessed by lack of motility) was
recorded.
2.3. Statistical analysis
Statistical analyses of toxicity data were performed
according to Newman (1995). Percentages of positive
response (embryogenesis success or survival) in each
vial (P ) were corrected by the mean response in
the control vials (Pc) according to Abbot�s formula,
which for positive responses takes the form: P 0 ¼100P=Pc.
For statistical analysis, the P 0 values were trans-
formed to the arcsine of the square root, and the values
were back-transformed for presentation.
Correlation analysis was performed according to Zar
(1984), and MDS followed Kruskal and Wish (1978).
All tests were performed using SPSS statistical soft-
ware.
3. Results
3.1. Environmental parameters
A profile of the environmental characteristics of the
sampling sites is reflected in Table 2. Despite high tem-
poral and microgeographical variability, some regula-
rities can be identified. The P1 and V2 sites are at
sampling time (low tide) under the influence of fresh
water inputs, while all other points show salinity higher
than 20 ppt. This is also the reason why no mussels are
present at P1. Seawater may be considered to be pol-
luted by organic matter when BOD5 > 4 mg/l. P1 and
particularly P2 and P3 exceeded this threshold. P1 and
P2 were also the sites with highest phosphate content, an
indicator of urban wastewater input (see Table 2).
Concerning the sediment, V2, B2 and particularly P3
where sandy sites, while all other samples consisted
mainly of silt. Consistently, P3 site showed positive
redox potential values, in contrast with all other sam-
ples, composed by finer sediment. The P1 site showed
the highest values of organic content and, along with V3
and P2, the lowest redox potentials.
3.2. Analytical chemistry
Table 3 summarizes the results from the chemical
analyses performed with the water, sediment and biota
Table 2
Physical and chemical parameters recorded in the water and sediment samples taken from the Galician Rias, and in the elutriates obtained from these sediments, in 1997 (first value)
and 1998 (second value)
Water Sediment Elutriate
T(�C)
S(ppt)
pH O2
(mg/l)
BOD5
(mg/l)
PO4
(lM)
Eh (mV) OM
(%)
63 lm(%)
T(�C)
S(ppt)
pH O2
(mg/l)
H2S
(mg/l)1 cm 3 cm 5 cm
V1 16.2 n.m. 8.0 8.7 1.01� 0.636 1.59 )83 )194 )311 3.3� 0.1 57.3 21.4 n.m. 7.0 4.0 n.m.
15.0 34.6 7.8 7.7 3.02� 0.867 1.10 )156 )212 )269 7.6� 0.60 46.4 22.1 32.5 7.2 6.2 <0.1
V2 15.7 n.m. 7.2 7.8 0.25� 0.233 1.09 )182 )313 )358 9.1� 1.4 12.2 21.2 n.m. 7.2 4.3 n.m.
15.5 6.3 7.7 9.7 2.61� 2.12 0.76 )70 )105 )167 4.59� 0.134 23.3 22.1 31.5 7.3 6.6 n.m.
V3 16.3 n.m. 7.3 8.4 0.35� 0.262 1.05 )278 )346 )356 2.23� 0.25 54.5 21.3 n.m. 6.7 3.6 n.m.
17.0 n.m. 7.7 8.7 0.65� 0.177 1.14 )249 )286 )311 5.31� 0.229 31.6 22.1 36.6 7.5 6.5 n.m.
P1 13.3 n.m. 6.8 8.7 3.29� 0.057 6.21 )352 )376 )393 12.2� 0.2 48.6 20.0 n.m. 7.1 0.0 2–2.5
13.8 17.0 7.1 6.4 4.57� 1.807 2.90 )215 )243 )278 11.6� 0.35 39.8 22.2 31.6 6.5 5.6 <0.1
P2 17.3 n.m. 7.3 5.4 9.50a 5.30 )240 )318 )347 4.9� 0.5 19.2 20.1 n.m. 7.4 1.5 <0.1
15.4 33.0 7.6 9.3 6.99� 0.194 7.55 )303 )329 )324 8.4� 0.34 24.7 22.2 32.3 7.2 1.1 <0.1
P3 17.1 n.m. 7.6 10.4 2.48� 0.340 1.97 109 98 86 2.6� 0.03 7.6 20.0 n.m. 7.7 4.5 <0.1
15.9 36.5 7.5 14.7 7.74� 3.682 0.90 134 137 78 0.69� 0.079 2.3 22.3 32.5 7.7 7.7 n.m.
A1 15.6 n.m. 8.2 12.0 2.07� 0.750 0.51 0.3 )136 )211 1.20� 0.191 15.9 21.8 n.m. 7.4 3.7 <0.1
16.1 24.8 8.5 3.00� 0.374 1.41 )131 )226 )325 7.73� 0.230 24.2 21.2 32.4 7.3 3.9 n.m.
A2 19.9 n.m. 8.1 12.0 3.41� 0.919 1.49 )295 )307 )325 2.10� 0.038 52.2 21.7 n.m. 7.0 3.8 <0.1
16.9 32.5 7.7 9.3 2.39� 0.572 1.02 )120 )220 )303 8.4� 0.37 65.9 21.4 31.9 7.8 4.1 n.m.
A3 19.5 n.m. 7.8 8.8 1.78� 0.113 0.74 )340 )362 )368 2.67� 0.27 29.1 21.8 n.m. 7.3 3.0 <0.1
15.2 31.9 7.4 7.7 1.82� 0.461 0.85 )92 )241 )334 9.29� 0.272 23.2 21.5 32.6 7.5 3.8 n.m.
B1 13.3 21.5 7.4 7.5 1.92� 0.382 1.60 )194 )227 )278 5.81� 0.155 63.2 20.9 32.4 7.7 5.1 n.m.
B2 16.7 30.7 7.5 10.2 2.63� 0.293 2.82 )67 )126 )279 2.67� 0.130 13.5 20.9 33.2 7.7 5.3 n.m.
B sites were sampled in 1998 only. T : temperature; S: salinity; O2: dissolved oxygen; BOD5: biological oxygen demand after 5 days (mean� s.d., n ¼ 3); PO4: phosphate (mean� s.d.,
n ¼ 3); Eh: redox potential at three depths (mean of three measurements per depth); OM: organic matter (mean� s.d., n ¼ 3). Annual maxima for BOD5 and PO4 are marked in
bold.aOriginal samples not measurable, extra sample taken on 3/10/97, n ¼ 1.
1214
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1209–1224
Table 3
Concentration of pollutants in the environmental samples taken from the Galician Rias in 1997 (first value) and 1998 (second value)
Water Sediment Mussels
Metals (nM) Metals (mg/kg dw) (lg/kg dw) Metals (mg/kg dw) Organics (lg/kg ww)
Cu Pb Cd Zn Cu Pb Cd ZnP
7PCB Cu Pb Cd ZnP
7PCB DDE HCB Aldrin
V1 20 25 0.24 60 43.1 73.5 0.20 164.5 4.2 7.15 6.39 0.478 249 6.1 0.53 <0.05 0.13
43.4 93.7 0.27 177.6 7.0 7.76 5.24 0.625 234 7.5 0.58 <0.05 0.14
V2 14 13 0.29 104 24.9 48.3 0.09 81.4 2.2 6.45 4.44 0.427 231 6.2 0.48 <0.05 0.16
28.1 53.1 0.07 136.4 4.5 6.92 4.55 0.831 284 7.2 0.59 <0.05 0.12
V3 27 36 0.20 105 236.1 173.7 0.35 196.9 16.2 8.74 8.79 0.547 291 8.3 0.50 <0.05 0.15
76.1 76.0 0.27 175.2 28.9 9.02 7.61 0.809 277 9.1 0.70 <0.05
P1 38 9 0.14 95 51.0 83.8 0.26 182.3 9.8 – – – – – – – –
48.4 75.0 0.31 197.7 18.6 – – – – – – – –
P2 1306 37 0.62 455 91.8 106.6 0.20 219.2 17.5 12.82 4.31 0.300 177 10 1.57 nd nd
206.0 99.3 0.24 393.8 31.9 8.39 2.72 0.400 218 9.8 1.39 <0.05 0.16
P3 3010 215 1.01 4819 3887 1423 3.45 17 609 155 27.43 11.68 0.487 362 14 3.07 nd nd
4282 2302 9.57 44 957 132 25.38 8.93 0.614 453 13 2.20 <0.05 0.20
A1 45 9 0.31 74 15.4 30.8 0.09 57.9 0.8 10.76 1.59 0.992 233 4.6 0.99 <0.05 0.28
32.3 32.1 0.13 82.0 1.4 11.25 1.73 0.814 171 5.3 0.87 <0.05 0.21
A2 28 11 0.28 51 51.3 47.8 0.39 136.5 2.3 9.89 2.99 0.891 312 2.6 0.74 <0.05 0.19
78.6 48.8 0.55 180.5 4.4 11.41 3.24 1.045 316 3.4 1.02 <0.05 0.17
A3 27 12 0.21 36 89.9 128.6 0.35 178.4 16.4 11.34 4.78 0.500 259 4.8 1.15 <0.05 0.25
90.9 143.1 0.40 217.7 12.9 12.53 5.00 0.701 304 5.4 0.89 <0.05 0.18
B1 – – – – – – – – – – – – – – – – –
95 28 1.74 992 201.0 161.4 2.00 1053.7 46.2 14.31 17.49 0.617 467 33 1.70 0.071 0.27
B2 – – – – – – – – – – – – – – – – –
75 32 0.51 479 15.1 40.4 0.09 88.8 27.1 11.55 8.24 0.584 447 21 0.95 <0.05 0.24
C 18 26 0.15 82
B sites were sampled in 1998 only. Annual maxima are marked in bold. dw: dry weight; ww: wet weight.P
7PCBs: sum of seven individual congeners (see text).
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1216 R. Beiras et al. / Chemosphere 52 (2003) 1209–1224
collected at the different sites two consecutive years. In
1997, P3 site showed maximum values for all pollutants
in all environmental matrices analyzed except for cad-
mium in the mussels (maximum in Ria of Arousa,
probably due to upwelling conditions) and Hg in sedi-
ment and mussels (maximum in P2 nearby). In 1998 two
additional sites were included in the monitoring. For
sediments P3 showed again the highest values, but B1
replaced P3 in maximum contents for Pb, Zn and or-
ganics in the mussels. Comparing the concentrations
showed in Table 2 with the levels usual in the coastal
environment (see Section 4), two groups of sampling
sites could be identified. The P3, P2, B1, and V3 sites can
be considered as chemically polluted, whereas the re-
maining sites showed in general low concentrations of
pollutants. The concentrations of Cu, Pb, Cd, Zn and
PCBs in sediments from P3 are particularly remarkable.
Logarithmic transformation normalized most toxi-
cant concentration data, except for Cu, Pb, Cd and Zn
in sediments, and Cu and Zn in water, due to the ex-
tremely high values at site P3. Non-metric correlation
(Spearman) was additionally performed when these data
were involved. According to the correlations (Table 4),
three distinct groups of pollutants could be identified: (i)
metals other than Hg, (ii) Hg and PCBs, (iii) orga-
nochlorine pesticides.
Comparing matrices, the contents of Hg, Cu, Pb, and
PCBs showed significant correlations between mussels
and sediment, and mussels and seawater (see Table 4).
The correlations between sediment and seawater though
were generally poor. For these pollutants, mussels seem
a suitable monitoring species. In contrast Cd showed
very low or even negative correlations between different
matrices.
3.3. Embryo-larval bioassays
Fig. 2 shows the embryogenesis success in sea-urchin
fertilized eggs exposed to sediment elutriates from the
sites sampled in 1998. The results of the sea-urchin
bioassays conducted with the 1997 sediments (not
shown) were fully consistent and showed the same pat-
tern. All sites from the Rias of Vigo (Fig. 2(a)), Arousa
(Fig. 2(c)), and O Burgo (Fig. 2(d)) showed no toxicity
to sea-urchin embryos. In contrast, P1 from Ria of
Pontevedra (Fig. 2(b), diamonds) showed moderate
toxicity, and P3 (triangles) showed extreme toxicity,
completely inhibiting embryogenesis even after 10 times
dilution in control seawater. Fig. 3 shows the same data
obtained in 1998 with mussels. Again sediment elutriates
from Ria of Vigo showed no embryotoxicity, while
elutriates from site P1 were moderately toxic and those
from P3 extremely toxic. For this species P2, A1 and A3
also showed moderate toxicity, inhibiting embryogenesis
when undiluted. Fig. 4 illustrates the data obtained for
ascidian bioasays in 1998. In full agreement with sea-
urchin data, sediment elutriates from all sites of Vigo,
Arousa and O Burgo were non-toxic, while sites from
Pontevedra were toxic. Again P3 showed higher toxicity
than the other sites of this Ria, although in this case the
highest dilution of the elutriate did suppress toxicity.
The toxicity of the sediment elutriates from the Pont-
evedra sites was also tested on prawn zoea larvae and
the data obtained are reflected in Fig. 5. Although these
larvae showed lower sensitivity than sea-urchin, mussel
and ascidian embryos, the pattern of response was
similar, with P3 being the most toxic site, and in 1997
elutriates from P1 and P2 showing toxicity when undi-
luted.
4. Discussion
Despite the high socioeconomical value of the Gali-
cian Rias derived from extensive aquaculture and coastal
fisheries, comprehensive studies concerning chemical
pollution in these ecosystems are scarce. Concerning
seawater, only Guerrero P�eerez et al. (1988) published
data for some metals, thus we have sampled in this study
a control site of oceanic characteristics (see C site in
Table 3). Based on these references, seawater from sites
P3, P2, B2 and B1 (in this order) showed marked metal
pollution, while the other sites are comparable to
background levels, except for Cu in A1 and P1.
Trace metal pollution in coastal sediments from
Galician Rias is more documented (Barreiro et al., 1988;
Nombela et al., 1994; Rubio et al., 1995, 1996, 2000;
Belzunce-Segarra et al., 1997a,b; Besada et al., 1997b,
1999; Carballeira et al., 1997). Despite the method-
ological variability in sample fractionation, digestion
and analysis, a consistent overall picture emerges from
the inspection of the data available for the geographical
area covered in the present study. Metal contents are
generally close to background levels except for localized
areas. High levels of Pb have been measured in the inner
part of Ria of Vigo, Cu in particular locations of the Ria
of Pontevedra, and Zn in O Burgo. There is no signifi-
cant Cd pollution in the Galician r�ııas. The case of Hg
has been discussed elsewhere (Beiras et al., 2002).
Accumulation of trace metals in mussels from the
Galician coast have been reported by Besada et al.
(1997a, 2002); PROVIGO (2000, 2001, metal analyses
by Besada), and Carballeira et al. (1997). In general, the
values found here are comparable with those previously
reported by Besada and coworkers, and markedly lower
than those from Carballeira et al. (1997).
For PCB and organochlorine pesticides previous
data are fragmentary compared to the geographic range
covered in the present study, but there are remarkable
concordances. For sediments, the range of variation inP7PCBs reported by OSPAR Commission (2000) in
some Galician Rias was 1.0–37 lg/kg dw, very similar to
Table 4
Matrix of correlation coefficients among the concentrations of pollutants in the sediments (Sed.), mussels (Mus.) and seawater (Wat.), significance level, and sample size
Sed. Mus. Wat.
Hg Cu Cd Pb ZnP
7PCB Hg Cu Cd Pb ZnP
7PCB DDE HCB Aldrin Hg Cu Cd Pb Zn
Sed. Hg 0.500� 0.250 0.549� 0.556� 0.573�� 0.512� )0.004 )0.482� 0.478� 0.088 0.478� 0.174 0.372 0.025 0.215 0.291 308 0.246 0.300
0.025 0.287 0.012 0.011 0.008 0.030 0.988 0.043 0.045 0.727 0.045 0.490 0.156 0.929 0.406 0.385 0.357 0.465 0.370
20 20 20 20 20 18 18 18 18 18 18 18 16 15 17 11 11 11 11
Cu 0.793��� 0.880��� 0.545 0.680�� 0.486� 0.519� )0.342 0.441 0.329 0.401 0.558� 0.471 )0.039 0.386 0.364 0.146 0.573 0.364
0.000 0.000 0.083 0.001 0.041 0.027 0.165 0.067 0.182 0.099 0.016 0.066 0.889 0.126 0.272 0.669 0.066 0.272
20 20 11 20 18 18 18 18 18 18 18 16 15 17 11 11 11 11
Cd 0.685�� 0.737��� 0.512� 0.112 0.625�� 0.049 0.490� 0.632�� 0.150 0.514� 0.564� 0.221 0.216 0.318 0.099 0.290 0.189
0.001 0.000 0.021 0.659 0.006 0.848 0.039 0.005 0.552 0.029 0.023 0.429 0.405 0.341 0.771 0.386 0.578
20 20 20 18 18 18 18 18 18 18 16 15 17 11 11 11 11
Pb 0.655� 0.735��� 0.494� 0.445 )0.445 0.684�� 0.391 0.531� 0.420 0.218 )0.177 0.431 0.255 0.082 0.618� 0.418
0.029 0.000 0.037 0.064 0.064 0.002 0.108 0.023 0.083 0.418 0.528 0.084 0.450 0.811 0.043 0.201
11 20 18 18 18 18 18 18 18 16 15 17 11 11 11 11
Zn 0.791��� 0.536� 0.595��� )0.364 0.521� 0.418 0.529� 0.651�� 0.477 )0.041 0.393 0.545 0.269 0.655� 0.545
0.000 0.022 0.009 0.137 0.027 0.084 0.024 0.003 0.062 0.884 0.118 0.083 0.424 0.029 0.083
20 18 18 18 18 18 18 18 16 15 17 11 11 11 11
PCBs 0.694�� 0.647��� )0.429 0.769��� 0.575� 0.723�� 0.668�� 0.412 0.114 0.606�� 0.673� 0.598 0.789�� 0.709�
0.001 0.004 0.076 0.000 0.013 0.001 0.002 0.112 0.686 0.010 0.023 0.052 0.004 0.015
18 18 18 18 18 18 18 16 15 17 11 11 11 11
Mus. Hg 0.336 )0.574� 0.442 0.158 0.562� 0.471� 0.085 )0.309 0.573� 0.612 0.518 0.793�� 0.794��
0.173 0.013 0.066 0.531 0.015 0.049 0.754 0.263 0.026 0.060 0.125 0.006 0.006
18 18 18 18 18 18 16 15 15 10 10 10 10
Cu )0.041 0.373 0.507� 0.390 0.892��� 0.592� 0.611� 0.383 0.915��� 0.716� 0.718� 0.648�
0.871 0.128 0.032 0.110 0.000 0.016 0.016 0.158 0.000 0.020 0.019 0.043
18 18 18 18 18 16 15 15 10 10 10 10
Cd )0.308 0.211 )0.410 )0.222 0.027 0.198 )0.395 0.139 )0.161 )0.436 )0.1150.214 0.400 0.091 0.375 0.920 0.479 0.146 0.701 0.656 0.208 0.751
18 18 18 18 16 15 15 10 10 10 10
Pb 0.710�� 0.742��� 0.267 0.294 )0.012 0.710�� 0.273 0.599 0.680� 0.648�
0.001 0.000 0.284 0.270 0.967 0.003 0.446 0.067 0.031 0.043
18 18 18 16 15 15 10 10 10 10
Zn 0.510� 0.357 0.512� 0.303 0.623� 0.382 0.507 0.323 0.430
0.030 0.146 0.042 0.272 0.013 0.276 0.134 0.362 0.214
18 18 16 15 15 10 10 10 10
PCBs 0.465 0.465 0.242 0.814��� 0.564 0.821�� 0.591 0.903���
0.052 0.052 0.384 0.000 0.090 0.004 0.072 0.000
18 18 15 15 10 10 10 10
DDE 0.796��� 0.628� 0.477 0.830�� 0.785�� 0.601 0.479
0.000 0.012 0.072 0.003 0.007 0.066 0.162
16 15 15 10 10 10 10
HCB 0.657�� 0.348 0.667 0.820� 0.024 0.048
0.008 0.243 0.071 0.013 0.955 0.911
15 13 8 8 8 8
Aldrin 0.032 0.667 0.532 )0.293 0.119
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1217
Table
4(c
ontinued)
Sed
.M
us.
Wat.
Hg
Cu
Cd
Pb
Zn
P7PCB
Hg
Cu
Cd
Pb
Zn
P7PCB
DDE
HCB
Aldrin
Hg
Cu
Cd
Pb
Zn
0.918
0.071
0.175
0.481
0.779
13
88
88
Wat.
Hg
0.242
0.742��
0.749��
0.699�
0.473
0.009
0.008
0.017
11
11
11
11
Cu
0.697�
0.464
0.709�
0.017
0.151
0.015
11
11
11
Cd
0.626�
0.697�
0.039
0.017
11
11
Pb
0.745��
Zn
0.008
11
Spea
rman
correlation
coeffi
cien
tsare
shown
forco
rrelationsinvolvingnon-n
orm
allydistributed
variables,
and
Pea
rson
correlation
coeffi
cien
tsoth
erwise(see
text).
� p<
0:05;��p<
0:01;��� p
<0:001.
1218 R. Beiras et al. / Chemosphere 52 (2003) 1209–1224
the range found in this study (0.8–46 lg/kg dw, except
for the highly polluted site P3). Franco et al. (1984) and
Nunes et al. (1997) reported PCB levels in sediments
from inner Ria of Vigo within the 10–40 lg/kg dw range,
comparable to the values found here in V1,2 and 3. For
wild mussels from Ria of Vigo,P
7PCBs ranged, except
for a single highly polluted site, from 2 to 15 lg/kg ww
(Gonz�aalez-Quijano et al., 1997, 1999), which agrees with
the present study (see Table 3). Our finding of maximum
PCBs accumulation in mussels close to A Coru~nna (33 lg/kg ww) is also in line with previous data (46 lg/kg ww;
OSPAR Commission, 2000). Finally, instances of PCB
and DDT pollution in certain sites of Ria of Pontevedra
(P3) had also been previously reported (Fumega et al.,
1984).
Water quality standards are intended to protect the
aquatic ecosystems vis a vis threats to their beneficial
uses. One of the main uses of water in the Galician Rias
is the shellfish culture. According to the European leg-
islation (Anon., 1979) the levels of pollutants in these
waters should not exceed those causing deleterious ef-
fects on the adult mollusks or their larvae. The LC50
values impairing bivalve embryogenesis in marine biv-
alve species have been reviewed by His et al. (2000), and
the geometric means are 40 lg/l for Cu, 320 lg/l for Zn,
968 lg/l for Pb and 2219 lg/l for Cd. Cu content in P2
and P3 seawater are above these values. Zn content in
P3 seawater is also very close to average LC50. There-
fore, current levels of Cu and Zn pollution in these
particular sites pose a threat to this important economic
activity.
The main goal of the present study was to link sedi-
ment toxicity bioassays with analytical chemistry. Sedi-
ments are preferred to water as environmental matrix
for both chemical and biological monitoring because
pollutant concentrations in sediments are much higher
and less variable in time and space (e.g. Beiras et al.,
2002). A qualitative approach will be first discussed.
Long et al. (1995) published biologically meaningful
sediment quality guidelines based on compilation of
large toxicity databases for marine organisms. The ef-
fects range-median (ERM) and effects range-low (ERL)
levels correspond to the 50 and 10 percentile of the
distribution of effective concentrations. Pollutant con-
centrations above ERM indicate that adverse effects on
the marine fauna would often occur. In Southern Spain,
DelValls and Chapman (1998) have established for some
pollutants site-specific sediment quality values on the
basis of integrative environmental quality assessment.
The comparison of the levels of pollutants found in the
Galician Rias with ERM values stress the high level of
pollution found in site P3. The contents of Cu, Pb and
Zn in P3 sediments both years, and the Zn in B1, exceed
ERM. In particular Cu content in P3 was 15 times ERM
and Zn content was 40–100 times ERM. These metals
are thus very likely responsible for the high toxicity
Fig. 3. Embryogenesis success for mussel exposed to elutriates from sediments sampled in Vigo (a), Pontevedra (b) and Arousa (c) in
1998. Symbols indicate sites 1 (diamonds), 2 (squares) and 3 (triangles).
Fig. 4. Embryogenesis success for ascidia exposed to elutriates from sediments sampled in Vigo (a), Pontevedra (b) and Arousa (c) in
1998. Symbols indicate sites 1 (diamonds), 2 (squares) and 3 (triangles).
R. Beiras et al. / Chemosphere 52 (2003) 1209–1224 1219
shown by the sediment samples from this site. No other
sampling sites show levels of pollutants in the sediment
exceeding ERM. This is consistent with the toxicological
data obtained with marine invertebrate embryos and
larvae, where P3 is the only site showing high toxicity
disregarding dilution or sampling date. Therefore, this
study supports the use of ERM values when biological
effects are not directly tested and are to be predicted
from sediment chemistry. In contrast, ERL values
were too conservative, and they are exceeded in many
Fig. 5. Survival of prawn larvae exposed to elutriates from
sediments sampled in Pontevedra in 1997 (a) and 1998 (b).
Symbols indicate site 1 (diamonds), 2 (squares) and 3 (trian-
gles).
1220 R. Beiras et al. / Chemosphere 52 (2003) 1209–1224
instances with no biological effects on the marine in-
vertebrate embryos associated. ERL values may still be
valuable to predict effects on infaunal organisms living
in contact with the sediment.
P1 was the only other site showing toxicity on sea-
urchin embryos, though to a much lesser extent. In 1997
this toxicity may be attributable to hydrogen sulfide, a
natural component of highly reduced sediments whose
concentration greatly increases with time of storage in
anaerobic conditions. The concentration of H2S mea-
sured in the P1 elutriate used for sea-urchin bioassay
was above 2 mg/l, whilst toxic effects were observed by
Knezovich et al. (1996) above 0.12 mg/l. However, in
1998 H2S content was negligible and the elutriates still
caused a certain degree of toxicity when undiluted. We
will further discuss this case below.
Secondly, we may try to establish a posteriori a
quantitative link between sediment toxicity and sedi-
ment chemistry. This falls beyond the aim of the ex-
perimental design of the present study, and it is
attempted here only for discussion purpose.
We have previously reported additive effects of metal
combinations on sea-urchin embryos (Fern�aandez and
Beiras, 2001). The present data allow the comparison of
the actual toxicity of the elutriates recorded in the sea-
urchin bioassays with the theoretical toxicity expected
according to an additive model for all the metals mea-
sured. With this aim, we have calculated the theoretical
toxicity units (TU) present in the elutriates attributable
to each metal analyzed. IfP
TU > 1 impaired larval
development is expected. The fraction of the metal
content of the sediment that is mobilized into the elu-
triated water has been measured independently in more
recent experiments with sediments and elutriates from
Ria of Pontevedra by using ASV (Lorenzo et al., 2002).
The average mobilization factors (lg/l in elutriate di-
vided by mg/kg DW in sediment) ranged between 10 and
150� 10�3 for Hg, Cu, Zn, Cd and Pb. Metal concen-
trations in the elutriates of the present study (½M �), es-timated using those data, were transformed into TU
following the expression;
TU ¼ ½M �=CL50M
where CL50M is the median effective concentration of
the metal on sea-urchin embryos, taken from Nacci et al.
(1986) for Zn and Fern�aandez and Beiras (2001) for the
other metals. The summation of the toxic units was 3.9
for P3 in 1997, 6.9 for P3 in 1998 and lower than 0.3 for
all other sites and sampling dates. Therefore, the cal-
culations made on the basis of an additive model with
the chemical data available predict that only P3 should
show toxicity on the sea-urchin embryos. This prediction
is thoroughly supported by the toxicological data with a
single exception: undiluted elutriates from P1 in 1998
were also toxic. The very simple additive model used
above must be refined in the future on the basis of more
comprehensive chemical and toxicological data. For
instance, it does not take into account the effects of sa-
linity on metal toxicity. As mentioned above, the P1 site
is affected by a fluvial input, and its elutriate shows
comparatively low pH and salinity (see Table 2), con-
ditions that enhance Cu and Zn toxicity. The effects of
pH and acid-volatile sulfides need also to be taken into
account. Alternatively, non-measured organic pollutants
may be responsible for the toxicity of 1998 P1 sediments.
A pleasure boats marina is located a few hundred meters
away from this site.
When multiple and heterogeneous information is
available for each site, ordination of sampling sites ac-
cording to the degree of pollution may be performed by
using multivariate techniques like MDS. This method
produces a graphical representation based on the ma-
trices of similarities of each site with every other site for
all the variables measured. When various potential
pollutants are measured in a set of sampling sites, data
are likely to include variables oscillating within very
different ranges, variables non-normally distributed, and
variables with no biological effects that only bring noise
into the model. Non-linear analyses are advocated for
these cases (Landis et al., 1997). By using non-linear
MDS, we have obtained the spatial configuration of the
sampling sites, for each sampling year independently, on
the basis of (i) the analytical chemistry data (Fig. 6(a)
and (c)), and (ii) the toxicological data obtained in the
(a) (b)
(c) (d)
Fig. 6. Spatial configuration of the sampling sites on the basis of analytical chemistry (a, c) and toxicological data (b, d) by using non-
linear MDS. (a, b) and (c, d) show the configurations for 1997 and 1998 data respectively.
R. Beiras et al. / Chemosphere 52 (2003) 1209–1224 1221
embryo-larval bioassays (Fig. 6(b) and (d)). A marked
degree of concordance between both types of configu-
rations can be observed. In 1997 sampling station P3,
and to a lesser extent P2, are discriminated from the
others in both the chemistry-based configuration and
the toxicology configuration, whilst P1 is separated in
the toxicological configuration only. This may be due
to the high levels of H2S and anoxic conditions recorded
in this elutriate. In 1998 again station P3 is markedly
separated from the others according to both chemical
and toxicological data. The other stations segregated in
the toxicological configuration is P1, although in this
case O2 and H2S were within the range allowing normal
embryo development, indicating that additional chemi-
cals not measured were responsible for the toxicity
found. Further, in the configuration based on analytical
chemistry (Fig. 6(c)) the sampling stations are arranged
according to their degree of pollution, from less polluted
(A1, A2, V2) on the upper right corner to the most
polluted (B1, P2, V3) on the bottom.
The pattern of pollution described above is further
supported by the data on bioaccumulation by the wild
mussels. On the basis of the OSPAR Commission (2000)
background levels (upper limit), and considering high an
1222 R. Beiras et al. / Chemosphere 52 (2003) 1209–1224
enrichment factor (measured level/background level) >5
and very high a factor >10, mussels from inner Ria of
Pontevedra showed very high levels of Hg and instances
of high Pb and PCBs, those close to A Coru~nna very high
levels of Pb and PCBs and high Hg, and mussels from
inner Ria of Vigo high levels of Pb and Hg. Mussels
from Arousa did not show high levels of any of these
pollutants.
Concerning the choice of species for bioassays, a re-
markable concordance in the patterns of response
among test species from very distant phylogenic groups
can be observed, suggesting that the toxicants mobilized
from the sediment act through non-selective universal
mechanisms. This is consistent with pointing at Cu and
Zn as presumably responsible for the toxicity observed.
However, the additional presence of very specific toxi-
cants cannot be discarded, and the use of a multispecies
battery of bioassays is still recommended, including
representatives of the main taxonomical groups of sea
resources to be protected. Along with the widely used
sea-urchins and bivalves, ascidians may cover neuro-
toxic compounds specifically toxic to chordates (Bellas
et al., 2001). A marine arthropod is also needed to detect
the presence of highly selective toxicants such as insec-
ticides. However, the use of prawn larvae assayed here
by first time showed important limitations, namely
limited biological material causing low replication, and
time-consuming experimental set-up. Prawn larvae are
also known to show low sensitivity to copper (Mari~nno-Balsa et al., 2000). Adult amphipods may be advisable,
and they are also suitable to test the toxicity of whole
sediment. Finally, a photosynthetic organism would
provide the bioassay with sensitivity against herbicides,
and it will be included in the battery of test species in the
future.
Acknowledgements
Authors are grateful with all the staff who lend a
helpful hand over these five years of intense work. J.I.
Lorenzo made the BOD and organic matter measure-
ments; J.C. Mari~nno-Balsa conducted the prawn
bioassays, A. Cobelo analyzed metals in seawater, L.
Pombar, Z. Romero, F. Schultze, B. Cambeiro, A.
Garc�ııa, I. Alves and M. Cerqueira provided technical
assistance.
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