Descargar - V Simposio Internacional de Ciencias del Mar

(2016)
XVIII Seminario Ibérico de Química Marina
Alicante (España), 20-22 Julio 2016
2
(2016)
Esta publicación debe citarse como:
VALLE C, AGUILAR J, ARECHAVALA P, ASENSIO L, BLASCO J, CABRERA R, COBELO A, CORBÍ H, CRAVO A, DELA-OSSA JA, DEL-PILAR Y, FERNÁNDEZ V, FERNÁNDEZ Y, FERRERO LM, FORCADA A, FORJA J, GIMÉNEZ F,
GÓMEZ A, GONZÁLEZ JM, IZQUIERDO D, LEÓN V, MARCO C, MARTÍNEZ E, ORTEGA T, RAMOS A, RUBIO E,
SÁNCHEZ JL, SÁNCHEZ P, SANTANA M, SANZ C, TOLEDO K, ZUBCOFF JJ (Ed.). 2016. Libro de Resúmenes. XVIII
Seminario Ibérico de Química Marina. Universidad de Alicante, Alicante. 114 pp. ISBN: 978-84-16724-18-5.
Edita:
Universidad de Alicante
Departamento de Ciencias del Mar y Biología Aplicada
ISBN:
978-84-16724-18-5
Diseño de portada:
Imagén de portada:
Maquetación:
Luis M. Ferrero
Pablo Arechavala
Luis M. Ferrero, Pablo Arechavala, Reme Cabrera
Esta obra está bajo una Licencia Creative Commons Atribución-NoComercial 4.0 Internacional
3
(2016)
Comité organizador
Abelardo Gómez, Universidad de Cádiz
Alexandra Cravo, Universidade do Algarve
Antonio Cobelo, Instituto de Investigaciones Marinas, CSIC
Jesús Forja, Universidad de Cádiz
Julián Blasco, Instituto de Ciencias Marinas de Andalucía, CSIC
Magdalena Santana Casiano, Universidad Las Palmas de Gran Canaria
Rocío Ponce, Universidad de Cádiz
Teodora Ortega, Universidad de Cádiz
Victor León, Instituto Español de Oceanografía
Comité organizador local
Aitor Forcada, Universidad de Alicante
Alfonso Ramos, Universidad de Alicante
Candela Marco, Universidad de Alicante
Carlos Sanz, Universidad de Alicante
Carlos Valle, Universidad de Alicante
David Izquierdo, Universidad de Alicante
Elena Martínez, Universidad de Alicante
Esther Rubio, Universidad de Alicante
Francisca Giménez, Universidad de Alicante
Hugo Corbí, Universidad de Alicante
Javier Aguilar, Universidad de Alicante
José Antonio de la Ossa, Universidad de Alicante
José Jacobo Zubcoff, Universidad de Alicante
José Luis Sánchez, Universidad de Alicante
José Miguel González, Universidad de Alicante
Kilian Toledo, Universidad de Alicante
Leticia Asensio, Universidad de Alicante
Luis M. Ferrero, Universidad de Alicante
Pablo Arechavala, Universidad de Alicante
4
(2016)
Pablo Sánchez, Universidad de Alicante
Reme Cabrera, Universidad de Cádiz
Victoria Fernández, Universidad de Alicante
Yoana Del Pilar, Universidad de Alicante
Yolanda Fernández, Universidad de Alicante
Comité científico
Xosé Anton Álvarez Salgado, Instituto de Investigaciones Marinas, CSIC
Abelardo Gómez Parra, Universidad de Cádiz
Antonio Tovar, Instituto de Ciencias Marinas de Andalucía, CSIC
Fiz F. Pérez, Instituto de Investigaciones Marinas, CSIC
Guillermo Grindlay, Universidad de Alicante
Isidro M. Pastor, Universidad de Alicante
Josep M. Gasol, Institut de Ciències del Mar, CSIC
Luís Lubián, Instituto de Ciencias Marinas de Andalucía, CSIC
María J. Bebianno, Universidade do Algarve
Melchor González Dávila, Universidad Las Palmas de Gran Canaria
Miguel Caetano, Instituto de Investigacao das Pescas e do Mar
Ricardo Prego, Instituto de Investigaciones Marinas, CSIC
5
(2016)
Instituciones organizadoras:
Universidad de Alicante
Universitat de Barcelona
Universidad de Cadiz
Universidad Católica de Valencia San Vicente Mártir
Universidad de Las Palmas de Gran Canaria
Universidad de Vigo
6
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AUTOR / Área temática / Tipo de comunicación .......................................................................... # Página
ÁLVAREZ-IGLESIAS et al. / Contaminación y ecotoxicología marinas / Póster ............................................ 9
ÁLVAREZ-VÁZQUEZ et al. / Contaminación y ecotoxicología marinas / Póster ......................................... 12
ÁLVAREZ-VÁZQUEZ et al. / Procesos biogeoquímicos / Oral ..................................................................... 14
APARICIO-GONZÁLEZ et al. / Cambio climático / Poster ............................................................................ 17
BAENA-NOGUERAS et al. / Contaminación y ecotoxicología marinas / Oral .............................................. 19
BASALLOTE et al. / Contaminación y ecotoxicología marinas / Oral ......................................................... 21
BASALLOTE et al. / Contaminación y ecotoxicología marinas / Póster ...................................................... 23
BEBIANNO et al. / Conferencia invitada ...................................................................................................... 25
BIEL-MAESO et al. / Contaminación y ecotoxicología marinas / Oral........................................................ 26
BRITO et al. / Procesos biogeoquímicos / Póster ........................................................................................ 28
COBELO-GARCÍA et al. / Contaminación y ecotoxicología marinas / Oral ................................................. 30
CORREIA et al. / Oceanografía / Póster ....................................................................................................... 32
COZAR / Conferencia invitada .................................................................................................................... 35
DÍAZ et al. / Contaminación y ecotoxicología marinas / Póster ................................................................. 38
DURÁ et al. / Procesos biogeoquímicos / Póster......................................................................................... 40
GARCÍA-GUERRA et al. / Contaminación y ecotoxicología marinas / Oral ................................................. 42
GÓMEZ-PARRA et al. / Procesos biogeoquímicos / Oral ............................................................................. 44
GONZALEZ et al. / Procesos biogeoquímicos / Oral.................................................................................... 46
GONZÁLEZ-GARCÍA et al. / Oceanografía / Oral ........................................................................................ 48
GONZÁLEZ-ORTEGÓN et al. / Procesos biogeoquímicos / Póster................................................................ 50
GUERRA et al. / Procesos biogeoquímicos / Póster..................................................................................... 52
JEREZ-MARTEL et al. / Procesos biogeoquímicos / Oral ............................................................................ 54
JIMÉNEZ-LÓPEZ et al. / Cambio climático / Oral ........................................................................................ 56
JIMÉNEZ-LÓPEZ et al. / Cambio climático / Oral ........................................................................................ 58
LEÓN et al. / Contaminación y ecotoxicología marinas / Oral ................................................................... 60
LÓPEZ-SÁNCHEZ et al. / Oceanografía / Póster .......................................................................................... 63
MARTÍNEZ-PÉREZ et al. / Oceanografía / Oral ........................................................................................... 65
MIL-HOMENS et al. / Contaminación y ecotoxicología marinas / Póster.................................................... 67
MORENO-ANDRÉS et al. / Contaminación y ecotoxicología marinas / Oral ............................................... 70
MUÑOZ-LECHUGA et al. / Procesos biogeoquímicos / Póster ..................................................................... 73
ORTEGA et al. / Cambio climático / Oral .................................................................................................... 75
ORTEGA et al. / Procesos biogeoquímicos / Póster ..................................................................................... 77
PÉREZ-ALMEIDA et al. / Oceanografía / Oral .............................................................................................. 79
PINARDI et al. / Conferencia invitada ......................................................................................................... 81
PINTADO-HERRERA et al. / Contaminación y ecotoxicología marinas / Oral ............................................. 82
RAIMUNDO et al. / Contaminación y ecotoxicología marinas / Oral........................................................... 84
RIBEIRO et al. / Contaminación y ecotoxicología marinas / Póster ............................................................ 86
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RODRÍGUEZ-ROMERO et al. / Cambio climático / Póster ............................................................................ 89
ROSA et al. / Oceanografía / Oral ............................................................................................................... 91
SAMPER et al. / Contaminación y ecotoxicología marinas / Póster............................................................. 93
SÁNCHEZ-QUILES et al. / Contaminación y ecotoxicología marinas / Oral ................................................ 94
SANTANA-CASIANO / Contaminación y ecotoxicología marinas / Oral ..................................................... 96
SENDRA et al. / Contaminación y ecotoxicología marinas / Oral ............................................................... 99
SIERRA et al. / Cambio climático / Oral.................................................................................................... 102
TOVAR-SÁNCHEZ et al. / Contaminación y ecotoxicología marinas / Póster ............................................ 105
TOVAR-SÁNCHEZ et al. / Contaminación y ecotoxicología marinas / Póster ............................................ 107
TRAVERSO-SOTO et al. / Contaminación y ecotoxicología marinas / Póster............................................ 109
TROMBINI et al. / Conferencia invitada .................................................................................................... 111
8
(2016)
Timing of lead sources and bioavailability in sediments from San Simón
Bay (NW Spain): core scanners as complementary tools
Paula Álvarez-Iglesias1,2, Isabel Rodríguez-Germade1, Belén Rubio1, Daniel Rey1, Begoña
Quintana3 & Jorge Millos2
1
Grupo GEOMA, Dpto. de Geociencias Marinas y Ordenación del Territorio, Facultad de Ciencias, Universidade de Vigo,
Vigo 36310 Spain
2
Servicio de Seguridad Alimentaria y Desarrollo Sostenible, C.A.C.T.I., Universidade de Vigo, Vigo 36310 Spain
3
Laboratorio de Radiaciones Ionizantes, Dpto. de Física Fundamental, Universidad de Salamanca, Salamanca, 37008 Spain
ABSTRACT
San Simón Bay (inner Ría de Vigo), a well-known Pb polluted area, was selected for monitoring the historical
Pb pollution and its diagenetic evolution based on pore water and sediment analyses on two sediment cores
collected close to the main Pb input, a ceramic factory. The age-model was constructed by Pb-210 dating and
corroborated by the detection of temporal markers in different cores (Cs-137, Pb maxima inputs). Very high Pb
total contents were observed, more than half in the recoverable fraction. Pb stable isotope ratios confirmed the
ceramic factory as the main Pb input, even after its closure. Results obtained by conventional and high-resolution
XRF techniques (Itrax core scanner) were combined to monitor Pb pollution. The Itrax core scanner provided
detailed information on core sediment composition, which could affect radionuclide activities, and metal
variability.
INTRODUCTION
Coastal sedimentary environments are very sensitive to
anthropogenic activities and have been the target of
numerous pollution studies [1]. In particular, very high Pb
levels were detected in San Simón Bay (inner Ría de Vigo,
NW Spain), being the main source an ancient ceramic
factory located at the NE coast of the Bay [2-3]. Sediments
of these bay are characterized by high organic matter
contents that fuel early diagenetic processes, which affect
to metal speciation [3]. The main aim of this work is to
monitor the historical and diagenetic evolution of Pb in the
sedimentary record of San Simón Bay by applying a
multidisciplinary approach with different analytical
techniques including XRF core scanners (Itrax) as useful
screening tools that allow obtaining high resolution data.
MATERIAL AND METHODS
Two cores were collected in the intertidal area of San
Simón Bay, close to the ceramic factory, in October 2010:
SS10T01 (0.27 m) and its replicate SS10T03 (0.36 m).
Once in the laboratory, cores were half-split and pore
waters were collected from SS10T01. These pore waters
were divided in three aliquots: for determining sulfates (by
UV-VIS spectrometry), sulfides (UV-VIS) and trace
elements (by ICP-MS). Afterwards, a U-channel was
extracted from each core, and analyzed by Itrax core
scanner. This allows obtaining the optical and radiographic
images of the cores and their qualitative elemental
composition by XRF in a few hours (measuring conditions:
Mo tube at 30 kV and 55 mA, 10 s exposure time and 500
µm step size). Cores were subsampled each 1 cm. This
resolution was selected for analyzing SS10T03, while 2-3
cm was chosen for SS10T01. Grain size distribution was
analyzed in both cores by laser diffraction. Core SS10T03
was dedicated to water content determination and
radionuclide analyses by γ-ray spectrometry (in particular
214
Pb, 210Pb and 137Cs for dating). The other analyses were
performed on core SS10T01. Contents of total carbon
(TC), nitrogen (TN), sulfur (TS) and inorganic carbon
(TIC) were determined by Elemental Analysis. Total
contents of major and trace elements were also determined
by conventional XRF. The methods proposed for trace
elements by the NWRI (Canada) [4] were used to obtain
the bioavailable, recoverable and total fractions of Pb (and
Ag, Cr, Cu, Zn). Lead stable isotope ratios were
determined on those extracts by ICP-MC-MS. Different
mixture models were applied to these ratios to obtain the
total anthropogenic component, the isotopic signature of
the Pb anthropogenic inputs and the relative contribution of
each Pb source [3]. All analytical work was performed at
the CACTI of the University of Vigo, except that on
radionuclides, that was performed at the LRI of the
University of Salamanca.
RESULTS AND DISCUSSION
9
(2016)
The studied sediments were mostly sandy silts with TOC,
TN, TS and TIC contents that are characteristic of the study
area. Two facies could be defined in core SS10T01: the
upper one (0-13.5 cm) which comprised dark-colored
sediments with higher TOC and TS contents than the lower
one (13.5-27 cm), with light-colored sediments and higher
TIC contents. A similar facies distribution was observed in
core SS10T03 (two facies: 0-16.5 cm, and below), but with
an additional coarse-sand layer near its bottom (Fig. 1).
Sediments are oxic until 2.5 cm, and then, suboxic. No Pb
was detected in pore waters (<0.200 µg L-1), where sulfate
content decreased with depth and sulfide content were low
or non-detectable. Lead total concentration was extremely
high, being higher in the upper facies (∼4.9%, coincident
with higher TOC contents) than in the lower facies (∼2.1%)
of core SS10T01 (Fig. 1). Pb profiles obtained with the
Itrax and conventional techniques (ICP-OES and XRF)
were very similar, with the advantage of the former in less
time consumption and sample pretreatment. Similar high
Pb levels were observed for core SS10T03. The highresolution Itrax data shown a high Pb variability in both
cores (57,950±22,320 peak area in SS10T01 and
56,500±21,200 p.a. in SS10T03) and also a marked
subsurficial maxima (Fig. 1). This maxima is coincident
with relatively high Ag, As, Cu and Zn contents. This is
probably due to the use of Zn oxide in the ceramic, As in
its glaze, and Ag-thread in the decoration of some pieces.
a)
0.40
0.80
206
Pb (%)
AGS (mm)
0.00
1.20
0.0 2.0 4.0 6.0 8.0 10.0
1.150
Pb/
207
1.160
Pb
1.170
AGS
TOC
PbITRAX
Pbbioav
Pbrecov
PbICP
20
ACKNOWLEDGEMENTS
1.0 1.5 2.0 2.5 3.0 3.5
b)
0.00
0
40000 80000 120000
TOC (%)
Pb (p.a.)
AGS (mm)
Pb (p.a.)
0.40
0.80
glazing
10
gypsum
Depth (cm)
0
1.20
0
Pbxs (Bq kg-1)
210
40000 80000 120000
0
20
40
60
80
0
AGS
Water
PbITRAX
137Cs/Fe
210Pb
xs
10
Depth (cm)
that of total Pb (Fig. 1). Then, high quantities of Pb are
adsorbed onto sedimentary particles, forming insoluble
salts and organic complexes, which are environmentally
available. Profiles of Pb stable ratios for the bioavailable
and recoverable fractions and the total Pb content were
similar, with a marked minimum at 11 cm (Fig. 1). These
ratios were very similar to the characteristic ratios of
gypsum from the ceramic factory [3]. The anthropogenic
Pb represents 99.8-99.9% of total Pb, the main input being
the ceramic industry, with a small contribution from petrol
combustion (∼4%).
Radionuclide profiles showed a clear grain-size influence,
with a marked increase toward the bottom core (Fig. 1)
probably due to higher contents of organic matter and finegrained sediments. Due to this the 137Cs specific activity
profile was normalized by the Fe content. Three relative
maximum were detected (Fig. 1). The first two were
considered time-markers (1987 and 1963), the last
discarded due to the described matrix-effect. The age
model was established by applying the CRS-MV model to
the in excess 210Pb specific activity profile [2] and
sedimentation rates of ∼7.6 mm a-1 were obtained. When
sediment cores from former studies are compared to the
studied cores [2,5] it was detected a relative displacement
of some characteristic features in the 137Cs specific activity
profile and in the total Pb content profile that could be used
as time markers, and then, sedimentation rates of the same
order (∼5-7 mm a-1) are obtained. According to the
corroborated chronology the studied sediments spans the
last 60 years. The Pb total profile show the starting of the
ceramic industry around 1972, with maximum discharges
around the beginning of the 1990s, and the persistence of
very high Pb levels in 2010, in spite of the industry closure
in 2001.
20
This work was supported by the projects IPT-3100002010-17
and
GLC2010-16688
(MICINN)
and
10MMA312022PR (XUGA). I. Rodríguez-Germade was
funded by a FPU scholarship (MECD) and P. ÁlvarezIglesias by the Ángeles Alvariño program (XUGA).
REFERENCES
30
40
0
10
20
30
40
50
Water content (%)
0.0
0.4
0.8
1.2
1.6
(137Cs / Fe)*103
Fig. 1. Depth-wise profiles of: a) average grain size (AGS),
TOC, Pb concentration and 206Pb/207Pb in the bioavailable,
recoverable and total fractions (by ITRAX and ICP-OES)
in core SS10T01; and b) mean grain size, water content,
total Pb content (ITRAX), 210Pb in excess and 137Cs Fenormalized specific activities in core SS10T03.
Regarding the distribution of these high Pb levels detected,
18% was recovered in the bioavailable fraction and 56% in
the recoverable fraction –their respective profiles similar to
1 - Prego R., Cobelo-García, A, 2003. Twentieth century
overview of heavy metals in the Galician Rias (NW Iberian
Peninsula). Environ Pollut, 121: 425-452.
2 - Álvarez-Iglesias P, Quintana B, Rubio B, Pérez-Arlucea
M, 2007. Sedimentation rates and trace metal input history
in intertidal sediments from San Simón Bay (Ría de Vigo,
NW Spain) derived from 210Pb and 137Cs chronology. J
Environ Radioact 98: 229-250.
3 - Álvarez-Iglesias P, Rubio B, Millos J, 2012. Isotopic
identification of natural vs. anthropogenic lead sources in
marine sediments from the inner Ría de Vigo (NW Spain).
Sci Tot Environ 437: 22-35.
10
(2016)
4 - Catalogue of National Water Research Institute, 2006.
Certified Reference Materials & Quality Assurance
Services. Canada Centre for Inland Waters National
Laboratory for Environmental Testing Burlington, Ontario,
Canada, v 5.7.
11
(2016)
20th century overview of industrial impact through trace elements content in
sediments from the Ria of Ferrol (NW Iberian Peninsula)
M.A. Álvarez-Vázquez1,2, R. Prego1, P. Álvarez-Iglesias3, M.C. Pedrosa-García4, S. Calvo1,
E. De Uña-Álvarez2, B. Quintana4, C. Vale5,6, M. Caetano5,6
1
Instituto de Investigaciones Marinas (IIM-CSIC), 36208 Vigo, Spain.
Grupo GEAAT, Área de Geografía Física (University of Vigo), 32004 Ourense, Spain.
3
Grupo GEOMA, Faculty of Marine Sciences (University of Vigo), 36310, Vigo, Spain.
4
Laboratorio de Radiaciones Ionizantes (LRI), Faculty of Sciences (University of Salamanca), 37008 Salamanca, Spain.
5
Instituto Português do Mar e da Atmosfera (IPMA), 1449-006 Lisboa, Portugal.
6
Interdisciplinary Centre for Marine and Environmental Research (CIIMAR/CIMAR), 4050-123 Porto, Portugal.
2
ABSTRACT
The Ría of Ferrol (Galicia, Spain) has hosted an important shipbuilding activity since the mid-18th century. The
release of trace elements to the system is closely related to the history and changes in the industrial sector. To
make an assessment of the changing impact, a short sediment core (50 cm depth) was recovered in the intertidal
estuarine area of the Grande-de-Xubia River, the main stream draining into the ria. Layers of 2 cm thickness of
fine sediments were separated and dried. Contents Al (normalizer element), Cd, Co, Cr, and Cu were determined
in each layer. The core was also dated by 210Pb γ-emission. Preindustrial layers were not achieved so, in order to
set a baseline, a comparison was made with data from the nearby Ria of Ares (close to a natural state). By using
the metal-Al relationships of this second dataset, the enrichment factor (EF), indicative of human pressure, was
calculated for each sample. These elements presented different degrees of enrichment and contamination. The
increasing order during the 20th century was as follows: negligible contamination for Cr, moderate contamination
for Co, severe contamination for Cd, and heavy contamination for Cu.
INTRODUCTION
Sediments, accumulated in estuaries over the time, are a
record of the water quality in the time when materials
precipitated. These sediments have a natural component
with origin in the rocks and soils within the river drainage
basin mainly, also by the activity of living organisms.
Since the Industrial revolution it is especially important the
input of waste materials from human activities, which have
“severely” affected many estuaries [1] being now among
the most impacted environments in the World [2] of the
Anthropocene. Trace elements (TEs) are naturally present
in sediments but, nowadays, their contents were noticeably
altered by anthropogenic contributions especially in coastal
areas where the human activity is intense, so they can be
used as tracers of human impact, by themselves and, also
by their association with other kind of contaminants.
The study of TEs in sediment cores, coupled with some
geochronology technique, allows the reconstruction of the
historical relationship between humans and estuaries, being
critical to first establish natural background values [3] to
make an assessment of the magnitude and evolution of the
human impact.
This communication presents a quick view of the historical
human impact, during the 20th century, in one of the most
industrialized areas of Galicia (NW Iberian Peninsula),
namely the Ria of Ferrol. The shipyard industry, settled in
the shore of this ria in the 18th century, has evolved and
reached its peak in the mid-20th century.
Fig. 1. Location map. Hydrographic information from
Augas-de-Galicia Co. Basemaps ©2014 ESRI.
12
(2016)
MATERIAL AND METHODS
The sediment core was sampled using a hand-driven Gouge
Augers Sampler in July 2012 during low tide. The
sampling point was in the estuary of the Grande-de-Xubia
River (Ria of Ferrol), in an area without apparent
bioturbation. Cores were on site divided into 2 cm layers.
Muddy samples were oven-dried. After acid digestion, Al
was analysed by Flame Atomic Absorption Spectrometry
(Varian SpectrAA 220 FS) in the Marine Biogeochmistry
laboratory, IIM-CSIC (Spain); Cadmium, Co, Cr and Cu by
Mass Inductively Coupled Plasma (Thermo-Elemental Xseries) in the Environmental Oceanography laboratory,
IPMA (Portugal). The cores were also dated by 210Pb γemission in the Ionizing Radiation Laboratory (LRI),
University of Salamanca (Spain). All procedures during
sampling, processing and analysis were performed using
trace metal clean techniques, and analytical results checked
by the use of certificate reference materials.
RESULTS AND DISCUSSION
The TEs contents in the core ranged 0.39-1.26 mgCd·kg-1,
9.9-28.0 mgCo·kg-1, 71-127 mgCr·kg-1 and 72-284
mgCu·kg-1. The highest extremes of the four elements are
over the ranges for unpolluted marine sediments [4] and the
background for estuaries in the same regional area [5], but
the contents of Co are withing the ranges for Galician soils
over granitic rocks [6] (dominant lithology in the Grandede-Xubia River basin). In consequence, these TEs are
suspicious of being human-enriched but the local lithology
may play a critical role in determining the extent of the
anthropogenic influence, a local reference is needed.
Since the core did not reach a preindustrial layer, in order
to compare with local background values, data from the
nearby Ria of Ares was used for comparison, because this
ria drainage area is composed by a similar lithology and the
contents of TEs in the sediments of the Eume River estuary
(see Fig. 1) were previously set as close to a natural state
(unpublished data). In order to avoid possible differences in
particle size, Al was chosen as normalizer element. An
enrichment factor (EF) was calculated as the quotient
between the TE-Al relationship (TE content divided by the
Al content) in each sample of the core from Ferrol and the
average TE-Al relationship in the background reference
(Ares), according to the following equation:
EF = ([TE]/[Al])Ferrol/([TE]/[Al])Reference
The calculated EFs for each layer of the core, jointly with
the results of the 210Pb dating, which allows building the
timeline, are presented in Fig. 2. To evaluate de degree of
anthropogenic impact a contamination criteria [7] was
used, selected categories in increasing degree of
contamination are: EF<1 negligible, 1<EF>2 possible,
2<EF>3 moderate, 3<EF>6 severe, 6<EF>9 very severe
and EF>9 heavy.
Fig. 2. Time variation of EFs and contamination criteria.
The four elements presented the same pattern of
enrichment’s heights and troughs, being the highest
enrichments around the 1950s when the industry blossoms
with the implementation of great oil-tankers building. The
EFs also reflects the beginning of modern shipbuilding in
1908 and the industry-decay after the 60s, when the sector
fell into crisis. In the 90s the industry was reactivated and
EFs rose again. Currently (2012), Cr and Co did not
present contamination (EFs were 0.9 and 1.1 respectively),
contamination due to Cu was moderate (2.8), and Cu (3.4)
remains in the range of severe contamination.
ACKNOWLEDGEMENTS
This study from the project “Land-sea exchange of trace
metals and its importance for marine phytoplankton in an
upwelling coast”, ref. CTM2011-28792-C02, was financed
by MINECO (http://www.co.ieo.es/proyectos/mitofito/).
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3 - Álvarez-Iglesias P, Rubio B, & Pérez-Arlucea M, 2006.
Reliability of subtidal sediments as “geochemical
recorders” of pollution input: San Simón Bay (Ría de Vigo,
NW Spain). Estuar. Coas. Shelf. S. 70(3):507-521.
4 - Prego R, & Cobelo-Garcıa A, 2003. Twentieth century
overview of heavy metals in the Galician Rias (NW Iberian
Peninsula). Environ. Poll. 121(3):425-452.
5 - Carballeira A, Carral E, Puente X, & Villares R, 2000.
Regional-scale monitoring of coastal contamination.
Nutrients and heavy metals in estuarine sediments and
organisms on the coast of Galicia (northwest Spain). Int. J.
Environ. Pollut. 13(1-6):534-572.
6 - Macías-Vázquez F, & Calvo de Anta R, 2009. Niveles
Genéricos de referencia de metales pesados y otros
elementos traza en suelos de Galicia. Xunta de Galicia.
7 - Álvarez-Vázquez MA, Bendicho C, & Prego R, 2014.
Ultrasonic slurry sampling combined with total reflection
X-ray spectrometry for multi-elemental analysis of coastal
sediments in a ria system. Microchem. J. 112:172-180.
13
(2016)
River-ria fluxes of dissolved trace elements:
pristine versus anthropogenic disturbed contributions
Miguel Ángel Álvarez-Vázquez1,2, Ricardo Prego1, Miguel Caetano3, Elena De Uña-Álvarez2
1
Instituto de Investigaciones Marinas (CSIC). 36208 Vigo, Spain.
Área de Geografía Física, Grupo GEAAT, Campus de Ourense (UVigo). 32004 Ourense, Spain.
3
Instituto Português do Mar e da Atmosfera (IPMA). 1449-006 Lisboa, Portugal.
2
ABSTRACT
Little information is available about how the land-sea contributions of pristine small world-rivers may vary in
contaminated conditions. Therefore, a comparison of fluvial discharges of dissolved trace elements was
performed in the Ria of Cedeira (NW Iberian Peninsula). Along a hydrological year samples were taken in the
main streams draining into the ria: the pristine Das-Mestas, the Condomiñas affected by sewage discharges, and
the Forcadas having a reservoir for water supply on its course. Daily flow data were provided by “Augas-deGalicia” and EMAVISA. Concentrations of dissolved trace elements were determined by ICP-MS. Clean
techniques were used. Precision and accuracy was checked by the analysis of CRM. River-ria fluxes were
quantified by a ratio estimator method derived from the Beale’s ratio estimator. Obtained results distinguish
three main concentration groups in the three rivers: (a) Al, As, Cr, Fe and Pb presented higher ranges in the
Condomiñas and lower in the Forcadas; (b) Cd, Cu and Mn were higher in Condomiñas and Ferreiras,; (c) Co
and Ni in Condomiñas were over those in the other two rivers. Mo presented similar ranges but two high outliers
in the Condomiñas. Thus, the reservoir influence depicted a river-ria flux decrease while sewage contributions
increased the river flux of trace elements.
INTRODUCTION
The study of the interactions between land and ocean has
gained interest in recent decades. This growing interest
includes the recommendation of LOICZ program to
incorporate the study of river inputs in the assessment of
coastal dynamics [1]. These authors state that the
construction of dams, deforestation, rising rates of
urbanization and overexploitation of water resources have
caused a change in river flows and the matter transported
by fluvial water. It is also noteworthy the effort and the
large number of publications arising from the “Land-Ocean
Interaction Study” (LOIS) about rivers draining into the
North Sea from the U. K. [2]. More recently, the study of
the contributions of trace elements in the marine area is a
developing topic, recommended in the international
program GEOTRACES. Trace elements in water systems
have undergone some monographs and book chapters [3],
but it seems a controversial issue about the lack of reliable
data on a global scale, particularly in small rivers.
The aim of this communication is to perform a fluvial
transport comparison of dissolved trace elements (D-TEs)
affected by different and paradigmatic types of human
pressure. In this way, the Ria of Cedeira (NW Iberian
Peninsula; Fig. 1), which presents three drainage basins
with similar climate and land use characteristics, was
selected.
MATERIAL AND MHETODS
Samples of water were monthly taken from October 2011
to October 2012 in the Condomiñas River (fluvial basin of
26 km2) affected by sewage discharges, in the pristine DasMestas River (82 km2) and in the Forcadas River (64 km2)
having a reservoir for water supply on its course.
Temperature, pH and conductivity were in situ measured.
14
(2016)
Figure 1: Location map of the Ria of Cedeira, drainage
basins and fluvial nets of the three main rivers flowing into
it. Sampling points are showed by circles.
River water samples were filtered inside a class-100
laminar flow cabinet using 0.45 µm acid-clean
polycarbonate filters. Samples were acidified to pH<2 with
suprapur 30% HCl. All procedures during sampling,
handling and analysis followed trace-metal-clean
techniques. Trace metals in the dissolved phase (Al, As,
Cd, Co, Cr, Cu, Fe, Mn, Mo, Ni, Pb and Zn) were
determined by ICP-MS (Thermo-Elemental X-Series). Iron
was analyzed by ETAAS (Varian SpectrAA 220).
Precision and accuracy of the procedure was checked by
the analysis of certified reference materials (SLRS-4)
getting a good agreement between the determined and
certified value. Daily flow data of Das-Mestas River was
provided by “Augas-de-Galicia” (station 445), Forcadas
River from EMAFESA and Condomiñas River using its
watershed ratio with the Das-Mestas River.
Fluvial loads of dissolved trace elements of each river to
the Ria of Cedeira were estimated by a ratio estimator
method derived from the Beale’s ratio estimator [4]. The
ratio estimator method consists in a load calculation
applying a correction factor to minimize bias taking into
account all the days with flow data, representing the ratio
between mean measured loads and mean current flow [4].
RESULTS
The three studied rivers presented the same seasonal flow
pattern during the hydrological year Oct. 2011 – Sept.
2012. Annual average flow rates were 0.54 m3·s-1 in the
Condomiñas River, 1.68 m3·s-1 in the Das-Mestas River
and 0.91 m3·s-1 in the Forcadas River.
Average concentrations of D-TEs in the Das-Mestas River
were lower than 1 nM for Cd (0.013 nM), Pb (0.041 nM),
Co (0.57 nM) and Mo (0.57 nM); in the range between 1
and 10 nM for Cu (2.6 nM), Ni (5.5 nM) and Cr (7.2 nM),
from 10 to 100 nM for As (46 nM) and Mn (48 nM); and
higher than 100 nM for Al (570 nM) and Fe (1080 nM).
Zinc remained under the detection limit (<4.5 nM).
Dissolved average concentrations of D-TEs in the
Condomiñas River were lower than 1 nM for Cd (0.023
nM), Pb (0.14 nM) and Mo (0.40 nM); in the range
between 1 and 10 nM for Co (2.7 nM), Cu (4.6 nM) and Cr
(9.6 nM); from 10 to 100 nM for Zn (18 nM), Ni (54 nM)
and As (58 nM); and higher than 100 nM for Al (850 nM)
and Mn (1800 nM). Average concentration of D-TEs in the
Forcadas River were lower than 1 nM for Mo (<0.02 nM),
Pb (0.018 nM), Cd (0.037 nM) and Co (0.81 nM); in the
range between 1 and 10 nM for Cr (1.5 nM), Cu (4.9 nM)
and Ni (6.3 nM); from 10 to 100 nM for As (10 nM), Zn
(56 nM) and Mn (93 nM); and higher than 100 nM for Al
(280 nM) and Fe (1070 nM).
DISCUSSION
The average concentrations of D-TEs in the Das-Mestas
River is, except for As, within the wide order of magnitude
of natural concentrations in the World rivers [3]. In this
World context Al, As, Fe and Mn present a similar order,
while other metals studied are around 10 times lower than
the average of 27 most large world rivers [3]. In
consequence, this river can be used as reference level to
assess the possible impact of the Codomiñas River,
possibly affected by urban sewage, and the Forcadas River,
with a water impoundment in its basin. Table 1 presents the
estimated annual yield normalized by drainage surface area
(km2), an appropriate way to compare rivers of different
basin size [5].
Table 1. Annual yield of dissolved trace elements
normalized by surface basin area (mol·yr-1·km-2).
Condomiñas
Al
730±110
As
39±5
Cd 0.018±0.005
Co
2.0±0.5
Cr
8.2±1.2
Cu
3.5±0.4
Fe
1330±340
Mn
339±95
Mo 0.35±0.11
Ni
53±5
Pb 0.094±0.024
Zn
15±3
Das-Mestas
510±130
27±6
0.032±0.021
0.44±0.12
4.9±0.8
7.8±5.0
720±120
33±5
0.40±0.10
5.9±2.3
0.13±0.10
3.1±0.2
Forcadas
182±108
4.7±2.4
0.021±0.015
0.37±0.20
0.55±0.36
3.0±1.7
600±310
95±69
0.011±0.006
3.1±1.7
0.0075±0.0040
53±50
The
anthropogenic
influence
promotes
higher
concentrations of several D-TEs. So, the urban influences
reflect a significant increase in Fe, Mn and Zn loads of the
Condomiñas River, while increases of Co, Cr and Ni may
be Non
due to local lythology [6]. Processes of
flocculation and precipitation within the water reservoir
presumably resulted in a depletion of D-TEs in the
Ferreiras River. In this river, the decrease of D-TEs fluxes
was important for Al, As, Cr, Cu, Mo and Pb.
ACKNOWLEDGEMENTS
This study from the project “Land-sea exchange of trace
metals and its importance for marine phytoplankton in an
upwelling coast”, ref. CTM2011-28792-C02, was financed
by MINECO (http://www.co.ieo.es/proyectos/mitofito/).
REFERENCES
1 - Syvitski JP, Vörösmarty CJ, Kettner AJ & Green P,
2005. Impact of humans on the flux of terrestrial sediment
to the global coastal ocean. Science, 308:376-380.
2 - Neal C & Davies H, 2003. Water quality fluxes for
eastern UK rivers entering the North Sea: a summary of
information from the Land Ocean Interaction Study. Sci
Total Environ, 314:821-882.
3 - Gaillardet J, Viers J & Dupré, B, 2003. Trace Elements
in River Waters. In: Turekian, K.K., Holland, H.D. (Eds.),
Treatise on Geochemistry. Elsevier, pp. 225-272.
4 - Joo M, Raymond MA, McNeil VH, Huggins R, Turner
RD & Choy, S, 2012. Estimates of sediment and nutrient
loads in 10 major catchments draining to the Great Barrier
Reef during 2006–2009. Mar Pollut Bull, 65:150-166.
15
(2016)
5 - Meybeck M, 2009. Fluvial export. Biogeochemistry of
Inland Waters. (Likens GE, Ed). Academis Press, 118-130.
6 - Prego R, Caetano M, Ospina-Alvarez N, Raimundo J, &
Vale C, 2014. Basin-scale contributions of Cr, Ni and Co
from Ortegal Complex to the surrounding coastal
environment. Sci Total Environ, 468:495-504.
16
(2016)
Flujo de CO2 en la interfase aire-agua de mar durante la campaña
RADMED_0216
Alberto Aparicio-González1, Safo Piñeiro1, Mari Carmen García-Martínez2, Marta Alvarez3,
Rosa Balbín1, Jose Luís López-Jurado1, Francina Moya2, Rocío Santiago1
1
Instituto Español de Oceanografía. Centro Oceanográfico de Baleares
Instituto Español de Oceanografía. Centro Oceanográfico de Málaga
3
Instituto Español de Oceanografía. Centro Oceanográfico de A Coruña
2
RESUMEN
Se muestran los resultados del flujo de CO2 en la interfase aire-agua de mar durante la campaña RADMED_0216
desarrollada en el Mediterráneo Occidental utilizando un medidor de pCO2 en continuo y su relación con otras
variables oceanográficas.
INTRODUCCIÓN
El Mediterráneo Occidental es un océano en
microescala, en el se desarrollan una gran variedad
de procesos físicos que también ocurren a nivel
global. En esta campaña se muestrean tanto zonas de
mar abierto como costeras incluyendo el delta del rio
Ebro, en las que podemos encontrar diversidad de
fenómenos. En general, se considera que las zonas
costeras son un sumidero de CO2 (dióxido de
carbono) (Borges, 2011; Cai, 2011) aunque todavía
no están descritos los mecanismos por los qué en
algunas ocasiones los sistemas costeros actúan como
fuentes de CO2 atmosférico y otras como sumidero
(Dai et al, 2013). Lee et al (2011) consideran que el
Mediterráneo Occidental es un importante sumidero
de CO2 atmosférico.
MATERIAL Y MÉTODOS
Dentro del proyecto RADMED se está muestreando
cada 3 meses la costa mediterránea española. En
RADMED_0216 (Febrero del 2016) se han realizado
un total de 81 estaciones de muestreo de CTD
dispuestas en radiales perpendiculares a la costa (Fig.
1). En la campaña desarrollada a bordo del B/O F.P.
Navarro se han realizado medidas en continuo de
pCO2 (presión parcial de CO2) en agua de mar de
superficie y en atmósfera, SST (temperatura
superficial),
SSS
(salinidad
superficial),
fluorescencia, presión atmosférica, batimetría y
velocidad y dirección del viento. Los análisis de
pCO2 se han realizado con un instrumento
SUNDANS (Surface UNderway carbon Dioxide
partial pressure ANalySer) que incluye un LICOR
LI-7000. Se han calculado los flujos de CO2 entre el
mar y la atmósfera con las ecuaciones de
Wanninkhof (1992) para conocer si el mar está
actuando como fuente o sumidero de CO2 en esta
época del año.
Fig. 1. Estaciones
RADMED_0216
de
muestreo
de
CTD
en
RESULTADOS Y DISCUSIÓN
Los resultados preliminares indican que el flujo de
CO2 a lo largo de la campaña (Fig. 2) ha sido
ligeramente negativo (valores entre 0 y -7 mmol m-2
d-1), excepto en el transecto entre el norte de Menorca
y Barcelona, donde se han encontrado valores mucho
más negativos, de hasta -30 mmol m-2 d-1. Los
valores negativos de FCO2 (flujo de CO2) indican una
entrada neta de CO2 desde la atmósfera hacia el mar,
por lo que se puede decir que durante todo el
muestreo el mar se ha comportado de forma muy
17
(2016)
débil como sumidero excepto en la zona más
próxima al Golfo de León que ha sido un sumidero
mayor. Para conocer cuáles de las variables que
hemos medio siguiendo el recorrido del barco son las
que influyen en esta distribución, se realizará un
Análisis de Componentes Principales. Además, para
una mejor interpretación contamos con las medidas
realizadas en toda la columna de agua mediante los
CTDs ya que así podemos identificar si el flujo pCO2
en el agua del mar puede estar asociado a estructuras
oceanográficas que afectan la solubilidad del CO2,
(zonas de upwelling, de convección, etc.).
REFERENCIAS
1 - Borges, A. V. (2011), Present day carbon dioxide
fluxes in the coastal ocean and possible feedbacks
under global change, in Oceans and the Atmospheric
Carbon Content, edited by P. Duarte, and J. M.
Santana-Casiano, chap. 3, pp. 47–77, Springer
Science +Business Media B.V.
2 - Cai, W.-J. (2011), Estuarine and coastal ocean
carbon paradox: CO2 sinks or sites of terrestrial
carbon incineration?, Annu. Rev. Mar. Sci., 3, 123–
145.
3 - Dai, M., Z. Cao, X. Guo, W. Zhai, Z. Liu, Z. Yin,
Y. Xu, J. Gan, J. Hu, and C. Du (2013), Why are
some marginal seas sources of atmospheric CO2?,
Geophys.
Res.
Lett.,
40,
2154–2158,
doi:10.1002/grl.50390.
4 - Lee, K., C. L. Sabine, T. Tanhua, T.-W. Kim, R.
A. Feely, and H.-C. Kim (2011), Roles of marginal
seas in absorbing and storing fossil fuel CO2, Energy
Environ. Sci., 4, 1133–1146.
5 - Wanninkhof, R. 1992. Relationship between wind
speed and gas exchange over the ocean. J. Geophys.
Res. 97(C5), 7373–7382. DOI: 10.1029/92JC00188
Fig. 2. FCO2 siguiendo el recorrido del barco en
RADMED_0216
AGRADECIMIENTOS
Este trabajo ha sido parcialmente financiado por el
programa Action-MED y el proyecto del Plan
Nacional ATHAPOC.
18
(2016)
Degradation of pharmaceuticals and personal care products in surface
waters: photolysis, hydrolysis and biodegradation kinetics
Rosa María Baena-Nogueras1, Eduardo González-Mazo1 and Pablo A. Lara-Martín1
1
Departamento de Química-Física, Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz, Campus de Excelencia
Internacional del Mar (CEI·MAR), Campus de Río San Pedro s/n 11510 Puerto Real, Cádiz, Spain.
ABSTRACT
We have focused this research on comparing the degradation kinetics of a wide number (n = 33) of frequently
detected pharmaceuticals and personal care products (PPCPs) considering different types of water, pH and solar
irradiation. For those compounds that were susceptible of photodegradation, their rates (k) varied from 0.02 to
30.48 h-1 at pH 7, being the lowest for antihypertensive and psychiatric drugs (t1/2 >1000 h). Modification of the
pH turned into faster disappearance of most of the PPCPs (e.g., k = 0.072 and 0.066 h-1 for atenolol and
carbamazepine at pH 4, respectively). The strongest decay (k = 0.03-3.73 h-1) was observed for antibiotics (e.g.,
ciprofloxacin and sulfamethoxazole), which also were hydrolyzed in dark controls (up to 50-60% in 28 days). On
the other hand, biodegradation was generally enhanced by marine bacteria, such in the case of mefenamic acid,
caffeine and triclosan (k = 0.019, 0.01 and 0.04 h-1, respectively), and was faster for anionic surfactants.
Comparing both processes, hydrochlorothiazide was eliminated exclusively by irradiation (t1/2= 0.15-0.43 h), as
well as diclofenac (t1/2=0.14-0.17 h), both being not biodegradable. Salicylic acid and phenylbutazone were
efficiently photo (t1/2 < 3 h) and biodegraded (t1/2 = 116-158 h), whereas some compounds such as ibuprofen,
carbamazepine and atenolol had low degradation rates by any of the processes tested (t1/2 = 23-2310 h), making
then susceptible to persist in the aquatic media.
INTRODUCTION
Significant levels of xenobiotic compounds can be often
measured in both freshwater and marine coastal systems
adjacent to populated areas [1]. Once in the water column,
there are different processes that can affect the
concentrations of organic contaminants. Among these
processes, we are going to focus in the different ways that
chemicals can be degraded. The degradation processes are
heavily influenced by a combination of the molecular
structures of the xenobiotic compounds and several
environmental factors such as bacterial communities, pH,
temperature, salinity, and irradiance, among others. In the
case of many PPCPs, there are recent studies stating that
photodegradation (caused by natural light) is one of the
most important processes removing these chemicals from
surface waters. It also promotes other degradation reactions
such as hydrolysis. Aerobic biodegradation (usually
mediated by bacteria that use oxygen as electron acceptor)
is also relevant in the removal of most PCPPs in the water
column. Recent works on the microbial degradation of
selected pharmaceuticals show very different behavior
depending on the substance considered, from relatively
high degradation speeds for acetaminophen or fluoxetine
(t1/2 < 12 days) to persistence of carbamazepine,
sulfamethoxazole and trimethoprim (t1/2 >100 days) [2].
Overall, the present study aims to determine and compare
the degradation kinetics in water for a wide range of
pharmaceutical and personal care products (n = 33) that are
often detected in sewage-impacted aquatic systems. More
specifically, we have conducted a series of experiments at
environmentally relevant concentrations (1-100 µg L-1) to
measure the photodegradation and biodegradation rates of
selected chemicals and the effect of several environmental
factors such as different pH and salinities. The results
presented here allow us to classify the different
contaminants in several groups depending whether they can
be efficiently photodegraded/biodegraded or not in natural
waters.
MATERIALS AND METHODS
Photolysis experiments were carried out following the
OECD guidelines Nº 316 for phototransformation of
chemicals in water by direct photolysis. Irradiation was
provided by a Suntest CPS+ simulator (Madrid, Spain)
equipped with a xenon lamp which simulates natural
sunlight in a wavelength range of 300-800 nm. Irradiance
was maintained constant at 500 W m-2 during all the
experiments and the temperature was monitored. The
solution employed in the experiments consisted of 250 mL
of HPLC grade water spiked to 100 ng mL-1 of target
compounds and nineteen sampling times were established
over a total exposure time of 24 hours. In order to study the
effect of the aqueous pH, it was adjusted to 4, 7 and 9,
depending on the experiment, using buffer solutions.
19
(2016)
Aerobic biodegradation experiments were carried out
following the OECD guidelines Nº 306 to study
biodegradability in seawater (July, 1992) using the shake
flask method. Freshwater was collected from Arcos de la
Frontera water reservoir, situated in Cadiz province (SW
Spain), whereas seawater was taken at the end of the tidal
creek Sancti Petri Chanel in Cadiz bay. Incubation
temperature was 19ºC, the same measured in the sampling
areas, and aeration was provided by magnetic agitation
(350 rpm). Once samples were acclimated, they were
spiked to 1 µg L-1 of each pharmaceutical and personal care
product. Sampling was performed by sacrificing two
bottles per time at 7 different sampling times over the
course of the experiment (28 days).
Once measured, the decreasing concentration of target
compounds versus time in both types of experiments
(photodegradation and biodegradation) was adjusted to a
pseudo first-order kinetic model using the equation:
Ln (Ct/C0) = -k t
The methodology used for both the analysis and
determination of PhACs and personal care products in
aqueous samples was that described in Baena-Nogueras et
al. (2016) [3] using solid phase extraction (SPE) and liquid
chromatography coupled to tandem mass spectrometry
(UPLC-QqQ-MS/MS) respectively.
have half-lives of 159 and 128 h, respectively, in seawater,
whereas they are not degraded in freshwater (Fig. 1d).
Nevertheless, further investigation is encouraged to
elucidate the structure of possible photodegradation and
biodegradation products, as well as their persistence and
toxicity, for a better understanding of these processes and
the final fate of PhACs and personal care products in
aquatic systems.
RESULTS AND DISCUSSION
Rresults differed depending on the compound considered,
as can be observed in Figure 1 for some selected chemicals.
Aqueous pH played a significant role in the
photodegradation processes, as well as the sample origin
and salinity for biodegradation. Hydrolysis, on the other
hand, was negligible for most compounds and, when
observed, was usually slower than the other two
degradation processes. Overall, the most recalcitrant
compounds were the psychiatric drugs carbamazepine and
amitriptyline, and the lipid regulator gemfibrozil, all of
them showing no degradation at any conditions tested,
except for photolysis in acidic and basic media, where
substances were removed up to 75%. Caffeine was only
biodegraded in seawater and phototransformed at pH 9.
Regarding personal care products, triclosan was sensitive
to photodegradation (Fig. 1a) whereas the surfactant SAS
showed no degradation by sunlight even when the pH was
modified. However, it was the most readily biodegradable
compound tested, with a half-life close to 6 h (Fig. 1b).
Several different trends were observed within the group of
analgesics/anti-inflammatories, which are characterized by
diverse physicochemical properties. As an example, there
were compounds mostly affected by irradiation such as
ketoprofen or diclofenac (k = 30.48 and 4.04 h-1,
respectively) (Figs. 1c).
One of the major novelties of this work is that there is
scarce information available concerning the degradation
processes of pharmaceuticals in marine environments. As
an example, we have reported for the first time that the
antibiotic trimethoprim and the bronchodilator albuterol
Fig. 1. Comparative photodegradation and biodegradation
curves for: a) triclosan, b) SAS, c) diclofenac, and d)
trimethoprim.
ACKNOWLEDGMENTS
This work has been carried out within a regional research
project (RNM 6613) funded by Consejería de Innovación,
Ciencia y Empresa (Junta de Andalucía), who also
provided a FPI fellowship.
REFERENCES
1 – Heberer T, 2012. Occurrence, fate and removal of
pharmaceutical residues in the aquatic environment: a
review of recent research data. Toxicol. Lett., 131:5-17.
2 - Benotti, M.J. and Brownawell, B.J., 2007. Distributions
of Pharmaceuticals in an Urban Estuary during both dryand wet.weather conditions. Environ. Sci. Technol.,
41:5795-5802.
3 - Baena-Nogueras, R.M., 2016. Determination of
pharmaceuticals in coastal systems using solid phase
extraction (SPE) followed by ultra performance liquid
chromatography – tándem mass spectrometry (UPLCMS/MS). Curr. Anal. Chem., 12:1-19.
20
(2016)
Effects of simulated CO2 diffuse leakage from sub-seabed storage sites on
trace elements mobility, under hydrostatic pressure (30 atm)
M. Dolores Basallote1, Karen M. Hammer2, Anders J. Olsen3, Inmaculada Riba1, Murat V.
Ardelan4
1
Cátedra UNESCO/UNITWIN WiCop. Departamento de Química-Física, Facultad de Ciencias del Mar y Ambientales,
Universidad de Cádiz. Polígono Río San Pedro s/n, 11510 Puerto Real (Cádiz), Spain.
2
SINTEF Materials and Chemistry, Marine Environmental Technology, 7465 Trondheim, Norway
3
Norwegian University of Science and Technology, Department of Biology, 7491 Trondheim, Norway
4
Norwegian University of Science and Technology, Department of Chemistry, 7491 Trondheim, Norway
ABSTRACT
Storing supercritical CO2 at geological structures below the sub-seabed, at great water depths, has been considered
as an option for the reduction of the atmospheric CO2 emission. Two main sources of accidental escape of CO2
have been proposed: from the transport facilities and from the storage areas. However the environmental effects of
potential CO2 leaks from the storage sites are still poorly understood. To study the effects of diffusive CO2
seepages on trace elements mobility, marine sediment from the Trondheim Fiord, collected at 250 m depth, was
subjected to diffusive CO2 seepage, under pressurized conditions (30 atmospheres), using a 1.4 m3 titanium tank.
Solubility and distribution of trace metals in seawater and sediment pore-water were analyzed in DGT samplers
which were deployed in the water and the tested sediment during a 9 days-CO2-seepage experiment. The results
showed that CO2 leakage affected the solubility, particle reactivity, and transportation rates of the studied elements
in sediments.
INTRODUCTION
The North of Europe and United States are heading the
carbon capture and stored (CCS) technology, not only by
projects in operation but also by research performed. The
Sleipner and Snøhvit projects have stored about 14 and 2
millions tones of carbon, respectively, since the beginning
of the activity [1]. Among all the CCS site selection around
the world many of them has been selected at offshore areas,
at relatively near coastal zone distances. The potential
damages in the completion or closure of the reservoir could
lead to CO2 leakages that could concern the surrounding
environment. Small (but continuous) seepages of CO2,
overpassing the sediment barrier to the seabed seawater may
be sufficient to cause ecological effects. Additionally, the
solubility of CO2 in seawater increases linearly with
increasing hydrostatic pressure and reducing temperature
[2]. Thus, it is expected the release of CO2 from storage
sites may not behave in the same way at seafloor depth that
acidification process at ocean surfaces.
While the sediments act as a sink for metals disposal, this is
not the final repository of metals and it could acts as source
of contaminants [3]. The changes in the factors controlling
metals bound sediments (i.e. granulometry, redox potential,
pH, organic carbon content, and dissolved oxygen) could
lead to the transformations to the most available forms of
metals for the marine organisms.
The presented work intends to describe the mobilization of
trace elements and heavy metals from sediments under
simulated diffuse seepages from sub-seabed CO2 storage
sites.
MATERIALS & METHODS
A laboratory-scale pressurized titanium tank (1.4 m3) was
used to execute a short-term experiment (9 days), under
pressurized conditions (30 atm) [4]. The sediments from the
Trondheim Fjord, collected at approximately 250 m depth,
were placed into the tank and subjected to diffusive CO2
flow (Fig. 1). Several DGT units, both for sediment and
water were placed inside the tank to follow the
accumulation of DGT labile elements in the tank during the
CO2 seepage. The DGTs were later analyzed for element
content in the incoming and outflowing seawater of the
tank, to determine difference in concentrations of the
studied trace elements. The trace metals determinations in
eluted DGT fractions were carried out using an InductivelyCoupled Plasma Mass Spectrometry (HR-ICP-MS, Thermo
Finnigan Element).
RESULTS & DISCUSSION
The presented results indicate an increase in the solubility of
metals (Fe, Mn, Al, U, Cr, and Cu) because of CO2 seepage.
21
(2016)
Furthermore, the increase of concentrations of DGT labile
fractions of the studied elements confirms the mobilization
and transformation of these elements due to the pH decrease
associated with possible CO2 leakage. Nevertheless, these
results cannot explain the precise mechanisms responsible
for the determined increase of the DGT-labile elements.
Fig. 1. Box containing the sediments and the DGT devices
placed into the Karl Eric Titanium Tank.
REFERENCES
1 - Michael, K., A. Golab, V. Shulakova, J. Ennis-King, G.
Allinson, S. Sharma, and T. Aiken. 2010. Geological storage
of CO2 in saline aquifers—A review of the experience from
existing storage operations. International Journal of
Greenhouse Gas Control 4:659-667.
2 - Koornneef, J., M. Spruijt, M. Molag, A. Ramírez, W.
Turkenburg, and A. Faaij. 2010. Quantitative risk
assessment of CO2 transport by pipelines—A review of
uncertainties and their impacts. Journal of Hazardous
Materials 177:12-27.
3 - Blasco, J., T. Gomes, T. García-Barrera, A. RodríguezRomero, M. Gonzalez-Rey, F. Morán-Roldán, C. Trombini,
M. Miotk, J.L. Gómez-Ariza, and M. Joao Bebianno. 2010.
Trace metal concentrations in sediments from the southwest
of the Iberian Peninsula. Scientia Marina 74:99-106.
4 - Ardelan, M.V., K. Sundeng, G.A. Slinde, N.S. Gjøsund,
T. Nordtug, A.J. Olsen, E. Steinnes, and T.A. Torp. 2012.
Impacts of Possible CO2 Seepage from Sub-Seabed Storage
on Trace Elements Mobility and Bacterial Distribution at
Sediment-Water Interface. Energy Procedia 23:449-461.
ACKNOWLODGEMENTS
Thanks to Syverin Lierhagen from Dept of Chemistry,
NTNU, for running the ICP-MS-analysis and to Sindre A.
Pedersen from Dept of Biology, NTNU, for his help during
alkalinity titration.
22
(2016)
CO2 leakage simulation; effects of the decreasing pH and the increasing
dissolved metals to the fertilization and larval development of Paracentrotus
lividus
M. Dolores Basallote1,, Araceli Rodríguez-Romero2, Manoela R. De Orte1,a, José M Quiroga,
T3. Ángel DelValls1, Inmaculada Riba1
1
Cátedra UNESCO/UNITWIN WiCop. Departamento de Química-Física, Facultad de Ciencias del Mar y Ambientales,
Universidad de Cádiz. Polígono Río San Pedro s/n, 11510 Puerto Real (Cádiz), Spain.
2
Departamento de Ecología y Gestión Costera. Instituto de Ciencias Marinas de Andalucía (CSIC). Campus Río San Pedro.
11510 Puerto Real (Cádiz). Spain.
3
Departamento de Tecnología de Medio Ambiente, Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz.
Polígono Río San Pedro s/n, 11510 Puerto Real (Cádiz). Spain.
ABSTRACT
Carbon capture and storage has become a new mitigation option to reduce anthropogenic CO2 emissions. The
effects of the CO2 related acidification, associated with unpredictable leaks of CO2 during the operational phases
were studied using the Paracentrotus lividus sea urchin liquid phase assays (Fertilization and embryo-larval
development tests). The urchin larvae exposed to elutriate of sediments with different metals concentration, which
were subjected to various pH treatments resulted in median toxic effect pH ranged from 6.33±0.02 and 6.91±0.01
for the egg fertilization, and between 6.66±0.03 and 7.16±0.01 for the larval development assays. The dissolved
metals concentration measured in the elutriates were associated with acidification. For all the sediment elutriates
tested the amount of the dissolved Zn increased in parallel with the pH reductions. Although correlated effects of
acidification, biological response and the presence of dissolved metals were observed from this work, further
research is required to properly explain the mechanisms associated with the increasing sediment toxicity because
of CO2 leakage.
INTRODUCTION
The Carbon Capture and Storage of CO2 (CCS) has the
potential to reduce the CO2 emissions from fossil fuel
combustion [1]. However, the CO2 must be stored for at
least hundreds of years if this technology intends to
contribute efficiently to reduce the atmospheric CO2
emissions. To date, 15 large-scale CCS projects are in
operation around the world with the capacity to capture up
to 28 million tones of CO2 per annum (Mtpa) [2].
The underground storage of CO2 is a relatively young
technology; therefore it remains a lack of understanding of
the behavior of the CO2 under leaks. The increase of H+
would be the effect of the mix and dissolution of CO2 into
seawater, which would lead to pH reductions, resulting in
seawater acidification processes.
Accumulating on the sediment surface by adsorption and
precipitation processes [3], the relevance of metals is
related to their toxicity, persistence and potential
bioaccumulation in marine organisms [4],[5]. Although
metals can be strongly bound to the sediments without
posing a threat to marine organisms, the remobilization of
the unconsolidated sediments, together with the expected
CO2-induced acidification as a consequence of leaks of
CO2 could cause important changes to the form in which
metals are present naturally in the environment increasing
their bioavailability to the marine organisms.
The aims of this work were to observe the direct effect of
decreasing pH, and whether CO2 related acidification could
affect the sediment metals behavior, increasing their
toxicity to the sea urchin larvae. To this end, the sea urchin
larvae were exposed to elutriates of sediments collected in
different littoral areas and subjected to various pH
treatments. Moreover, dissolved metal concentrations in
the sediment elutriates were measured, intending to
correlate sediment elutriates toxicity and pH reductions
with the biological responses.
MATERIALS & METHODS
Test sediments were collected from two littoral areas in the
Gulf of Cádiz, along the Southwestern Atlantic Coast of
Spain: the Bay of Cádiz and the Ría of Huelva. The
sediment sites were selected on the basis of the best
available information to represent presumably low and high
23
(2016)
levels of metals contamination [6]. Additionally, the subseafloor of the South West part of the Iberian Peninsula has
been selected as one of the possible CO2 storage sites in
Europe [7].
A laboratory-scale-CO2-injection system, designed to
conduct ecotoxicological assays was used to work with
sediment elutriates, employing a range of pH treatments
[8]. The sediment elutriation procedure was performed
according to modifications of the USEPA method (1998)
[9] and Environment Canada (1994) [10], in 2 L test
vessels. The pH treatments ranged from 8.0 to 6.0 (two
replicates per treatment) for each one of the sediment tests.
The fertilization tests were adapted from Ghirardini et al.
[11] and conducted in 25 mL polyethylene vessels
containing 10 mL of the studied sediment elutriates. The
embryo-larval development procedure was based on the
methods developed by Fernandez and Beiras [12].
The median effect concentration (EC50) was calculated for
the toxic effects associated with pH reductions using the
linear interpolation method. Statistical differences (p ˂
0.05, p ˂ 0.01) in fertilization and developmental success
between the sediment (MAZ, and ML) elutriates in
reference to the TRO sediment (considered the relatively
unpolluted sediment) were calculated. A multivariate
analysis was conducted using principal component analysis
(PCA) as the extraction procedure to describe the
distribution of the data with the minimum loss of
information.
RESULT AND DISCUSSION
According to the percentages of fines the sediment from
MAZ was classified as muddy sand, while the sediments
from TRO and ML sites were classified as sandy mud. The
sediments from the Ría of Huelva, MAZ and ML, exhibited
the highest metal concentrations.
As expected, the highest pCO2 were recorded at pH 6.0,
given that the highest amounts of CO2 were injected to
reach the lowest pH treatments. A similar pattern of CO32reduction as the pH decreased as well as the saturation
states of aragonite (ΩArag) and calcite (ΩCal) were shown in
all the elutriates tested. For the presented experiments, CO2
gas was injected in order to modify and control the pH in
the aquaria, then the balance between carbon species was
altered. Since CO2 concentration alters the TIC ([CO2],
[H2CO3], [HCO3-] and [CO32-]) in the system, the
bicarbonate ions also vary, leading to an increase of the
total alkalinity
The EC50 was estimated based on the fertilization failure
and the abnormal larval development, for the larvae
exposed to the sediment elutriates subjected to the pH
treatments. The EC50 ranged from 6.33±0.00 to 6.91±0.01
for the egg fertilization assay. The EC50 calculated for the
embryo-larval assay ranged between 6.66±0.03 and
7.16±0.01. According to our results, the acute sea urchin
larvae test (egg fertilization) and the chronic test (embryolarval development) are useful tests to study the effects of
CO2 induced acidification, with a slightly greater
sensitivity observed with regard to acidification in the
embryo-larval development assay.
Among the metals analyzed, Co, Zn, As, Cu, and Fe
exhibited detectable concentrations in the sediment
elutriates, passing from the sediment into the liquid phase.
The dissolved metals may be easily available to aquatic
organisms and therefore they could present more toxicity
than particulate metals.
ACKNOWLEDGMENTS
The Spanish Ministerio de Economía y Competitividad
partially supported the research work presented in this
document, under grant reference CTM2011-28437-C0202/TECNO and CTM2012-36476-C02-01/TECNO. The
authors are grateful to international Grant from Bank
Santander/UNESCO Chair UNITWIN/WiCop for funding
this work.
REFERENCES
1 – IPCC, 2005. Cambridge University Press, Cambridge, New
York
2 - Global CCS Institute, 2015. Global CCS Institute, Canberra,
Australia. (Global CCS Institute 2015)
3 - Atkinson, C.A., Jolley, D.F. and Simpson, S.L., 2007. Effect of
overlying water pH, dissolved oxygen, salinity and sediment
disturbances on metal release and sequestration from metal
contaminated marine sediments. Chemosphere 69(9), 1428-1437
4 - Tam N, Wong Y,2000. Spatial variation of heavy metals in
surface sediments of Hong Kong mangrove swamps.
Environmental Pollution 110 (2):195-205
5 - Allen HE., 1993. The significance of trace metal speciation for
water, sediment and soil quality criteria and standards. Science of
The Total Environment 134, Supplement 1:23-45
6 - Riba I, Forja JM, Gómez-Parra A, DelValls TÁ, 2004.
Sediment quality in littoral regions of the Gulf of Cádiz: a triad
approach to address the influence of mining activities.
Environmental Pollution 132 (2):341-353
7 – GeoCapacity, 2009. Assessing European Capacity for
Geological Storage of Carbon Dioxide. D16. WP2 Report Storage
Capacity
8 - Basallote M, Rodríguez-Romero A, Blasco J, DelValls A, Riba
I, 2012. Lethal effects on different marine organisms, associated
with sediment–seawater acidification deriving from CO2 leakage.
Environmental Science and Pollution Research 19 (7):2550-2560
9 – USEPA, 1998. Evaluation of Dredged Material Proposed For
Discharge in Waters of the U.S. - Testing Manual. vol
EPA/823/B/98-004. US Army Corps of Engineers. United State
Environmental Protection Agency
10 - Environment Canada, 1994. Guidance Document on
Collection and Preparation of Sediments for Physicochemical
Characterization and Biological Testing. Environmental Protetion
Services.
11 - Volpi Ghirardini A, Arizzi Novelli A, Tagliapietra D (2005)
Sediment toxicity assessment in the Lagoon of Venice (Italy) using
Paracentrotus lividus (Echinodermata: Echinoidea) fertilization
and embryo bioassays. Environment International 31 (7):10651077
12 - Fernández N, Beiras R, 2001. Combined Toxicity of
Dissolved Mercury With Copper, Lead and Cadmium on
Embryogenesis and Early Larval Growth of the Paracentrotus
Lividus Sea-Urchin
24
(2016)
Ecotoxic effects of Deep-Sea Mining
M. J. Bebianno, N.C. Mestre and the MIDAS consortium
CIMA, University of Algarve. Campus de Gambelas, 800-139 Faro, Portugal
ABSTRACT
In the last decades there was an increasing knowledge about deep-sea non-living resources and of their
commercial interest for being exploited namely polymetallic sulphides, manganese nodules, cobalt-rich
ferromanganese crusts, methane hydrates and rare earth elements. The future exploitation of deep-sea mineral
and energy resources will inevitably release toxic compounds that might have significant impact on deep-sea
fauna and on their biodiversity. For this reason there is an urgent need to identify the appropriate environmental
guidelines and develop the environmental practices to ensure that industry will exploit those resources under the
best environmental practices. Because the data currently available on the ecotoxicological risks of the potential
release of toxic mixtures from deep-sea mining is scarce the project MIDAS addresses these issues. In this
presentation results will be presented on ecotoxicological information relevant to be used in the future guidelines
for deep-sea mining. This work was developed under the MIDAS project, funded by the European Commission
7th Framework Programme under the theme “Sustainable management of Europe’s deep sea and sub-seafloor
resources” (Grant Agreement 603418).
25
(2016)
Presencia, distribución y riesgo ambiental de productos farmacéuticos en el
Golfo de Cádiz (SO, España)
Miriam Biel-Maeso1, Rosa María Baena-Nogueras1, Carmen Corada-Fernández1 y Pablo A.
Lara-Martín1
1
Departamento de Química-Física, Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz, Campus de Excelencia
Internacional del Mar (CEI·MAR), Campus de Río San Pedro s/n 11510 Puerto Real, Cádiz, Spain.
RESUMEN
Se han analizado más de 80 fármacos diferentes (incluyendo antiinflamatorios, reguladores de lípidos,
antihipertensivos, antidepresivos, antibióticos,…) en aguas del Golfo de Cádiz mediante extracción en fase sólida
(SPE) seguido de cromatografía líquida de ultra resolución – espectrometría de masas en tándem (LC-MS/MS).
Destaca por hallarse en mayor concentración sustancias antimicrobianas (>50%) (ej.: triclosán), seguidos de los
analgésicos y antiinflamatorios (>20%). Algunas familias como los antihipertensivos y reguladores lipídicos
fueron halladas en porcentajes muy bajos (<3%). Las mayores concentraciones de fármacos se midieron en el
interior de la Bahía de Cádiz (1000-4000 ng L-1) y las más bajas en el área oceánica (50-150 ng L-1), siendo el
transecto más contaminado el localizado en el estuario del Río Guadalete. En este área además, se ha observado la
existencia de un comportamiento conservativo para algunos compuestos persistentes tales como carbamazepina y
atenolol. Un análisis de riesgo medioambiental revela que no se prevén efectos negativos para la mayoría de los
compuestos detectados, a excepción de dos antibióticos de la familia de las fluoroquinolonas (ciprofloxacino y
ofloxacino) en la zona del Caño de Sancti Petri y una sulfonamida (sulfametoxazol) y el antimicrobiano (triclosán)
hallados en el estuario del Río Guadalete, presentando coeficientes de riesgo (HQ) entre 3.04 y 351.
INTRODUCCIÓN
El estudio del destino de los contaminantes emergentes en
el medio ambiente es crucial para realizar una buena
gestión de los recursos hídricos. Su calidad se ve
seriamente afectada por la creciente presión urbana y
agrícola. Entre los contaminantes emergentes presentes en
las aguas cabe destacar fármacos, productos de higiene y
cuidado personal, compuestos perfluorados y hormonas.
La principal fuente de entrada de estos compuestos en el
medio ambiente acuático son las aguas residuales, aunque
también cabe destacar el papel de la agricultura y ganadería
como fuentes de contaminación difusa de pesticidas y
antibióticos, respectivamente. En la mayoría de los casos
su eliminación en las estaciones depuradoras de agua
residual convencionales (E.D.A.R.) no es completa e
incluso puede afectar a la producción de agua potable [1].
El principal objetivo de este trabajo es llevar a cabo el
primer estudio sobre las fuentes, distribución y el impacto
de un gran número de productos farmacéuticos (> 80) en el
Golfo de Cádiz. Con el fin de evaluar el estado del medio
ambiente acuático de la zona se realizaron 6 transectos de
monitorización espacial entre las aguas costeras de la Bahía
de Cádiz que comprendió tres ríos afectados por las
descargas de efluentes procedentes de E.D.A.R. cercanas
(Río Guadalete, Caño de Sancti-Petri y Río San Pedro) y
tres transectos oceánicos que partieron desde Cabo
Trafalgar, Caño de Sancti-Petri y desembocadura del Río
Guadalquivir (Fig. 1). Posteriormente se realizó un análisis
del riesgo ambiental utilizando las concentraciones
determinadas y bibliografía disponible sobre la toxicidad
de fármacos en especies acuáticas.
Bahía de Cádiz
GOLFO DE CÁDIZ
Guadalquivir
Sancti-Petri
Cabo Trafalgar
BAHÍA DE CÁDIZ
Río Guadalete
Río San Pedro
Caño Sancti-Petri
Fig. 1. Área de muestreo en el Golfo de Cádiz.
26
(2016)
Las campañas se realizaron en verano de 2015 en 6
transectos localizados en el Golfo de Cádiz (Fig. 1). Las
muestras de agua fueron recogidas en botellas ámbar de 1L
previamente lavadas, filtradas por 0.45 micras y
almacenadas a 4ºC hasta su procesamiento en el
laboratorio. La metodología utilizada para el análisis de 83
compuestos farmacéuticos se realizó por la técnica de
extracción en fase sólida (SPE), seguida de cromatografía
líquida de ultra resolución conectado a un detector de
espectrometría de masas (UPLC-QqQ-MS/MS) [2].
Respecto al análisis ambiental, se consideraron las
concentraciones más altas como el peor escenario posible
(PECs) y aquellos valores límite a partir de los cuáles ya no
se producen efectos según estudios toxicológicos (PNECs).
Ambos valores permiten establecer un ratio PEC/PNEC
que se corresponde con el coeficiente de riesgo (HQ) y por
el cuál se define la existencia de cierto riesgo ecológico en
caso de encontrarse por encima de 1.
Para el área más contaminada, el estuario del Río
Guadalete, se observa que la relación entre la
concentración de algunos fármacos y la salinidad era
generalmente lineal (Fig. 4). Tal relación se debe a la
mezcla longitudinal de las aguas residuales procedentes de
la E.D.A.R. de Jerez de la Frontera que se van mezclando
con el agua salina a lo largo del recorrido del transecto y ve
mermada su concentración, surgiendo un comportamiento
conservativo como consecuencia de su baja capacidad de
adsorción y degradación [3]. Además, se hallaron algunos
de los mayores valores de riesgo ambiental (HQ), siendo
máximos para sulfametoxazol y triclosán.
Concentración [ng L-1]
MATERIAL Y MÉTODOS
Carbamazepina
40
y = -0.7523x + 27.203
R² = 0.9727
30
20
10
0
0
10
20
Salinidad (‰)
RESULTADOS Y DISCUSIÓN
Como puede observarse en la Fig. 2, los transectos con
mayor concentración por fármacos están localizados en el
interior de la Bahía de Cádiz, siendo el transecto del Río
Guadalete el que posee mayor contaminación con valores
que oscilan entre 1000-4000 ng L-1. Por otra parte, se
detectaron valores inferiores en los transectos oceánicos del
Golfo de Cádiz, encontrándose la media de éstos en un
rango de 50-150 ng L-1.
Log [PhACs] (ng L-1)
Primer cuartil
10000
Valor más bajo
Mediana
Valor más alto
y = -3.1955x + 116.62
R² = 0.9583
100
50
0
0
10
20
Salinidad (‰)
30
40
Fig. 4. Relación entre la concentración y la salinidad para
los productos farmacéuticos seleccionados.
AGRADECIMIENTOS
100
10
Transectos Bahía de Cadíz
Guadalete
Caño SP
Transectos oceánicos
San Pedro Guadalquivir Sancti-Petri
Trafalgar
Fig. 2. Diagramas de cajas que muestra la concentración
hallada en los 6 transectos de estudio.
Entre las familias de fármacos encontradas en el área de
estudio cabe destacar los antimicrobianos, seguido de
antiinflamatorios, antidepresivos y estimulantes (Fig. 3).
1%
21%
57%
40
Tercer cuartil
1000
1
Concentración [ng L-1]
Atenolol
150
30
2%
9%
10%
Analgésicos y antiinflamatorios
Antihipertensivos y reguladores lipídicos
Antidepresivos y estimulantes
Antibióticos
Antimicrobianos
Otros fármacos
Fig. 3. Porcentaje de las familias de fármacos encontradas
en el área de estudio.
Este trabajo ha sido llevado a cabo como parte del proyecto
RNM-6613 financiado por la Consejería de Innovación,
Ciencia y Empresa de la Junta de Andalucía y por el Plan
de Campaña Oceanográfica STOCA 201509 en el B/O
Angeles Alvariño coordinado por Instituto Español de
Oceanografía (IEO).
REFERENCIAS
1 - Heberer T, 2002. Occurrence, fate, and removal of
pharmaceutical residues in the aquatic environment.
Toxicol. Lett. 131 (1-2): 5-17.
2 - Baena-Nogueras RM, Pintado-Herrera MG, GonzálezMazo E. & Lara-Martín PA. 2015. Determination of
Pharmaceuticals in Coastal Systems Using Solid Phase
Extraction (SPE) Followed by Ultra Performance Liquid
Chromatography – tandem Mass Spectrometry (UPLCMS/MS). Curr. Anal. Chem., 11 (4): 1-19.
3 - Lara-Martín PA., González-Mazo E., Petrovic M.,
Barceló D., Brownawell BJ. 2014. Occurrence, distribution
and
partitioning
of
nonionic
surfactants
and
pharmaceuticals in the urbanized Long Island Sound
Estuary (NY). Mar. Pollut. Bull., 85 (2): 710-719.
27
(2016)
Distribution of rare earth elements in estuarine sediments from the Tagus
Estuary (Portugal): Evidence of anthropogenic contamination
Pedro Brito1,2, Isabel Caçador2, Ricardo Prego3, Mário Mil-Homens1, Miguel Caetano1
1
IPMA - Portuguese Institute of Sea and Atmosphere, Rua Alfredo Magalhães Ramalho, 6, 1495-006 Lisbon, Portugal
FCUL - Faculdade de Ciências, Universidade de Lisboa, Campo Grande, 1749-016 Lisbon, Portugal
3
CSIC-IIM - Marine Research Institute (CSIC), Av. Eduardo Cabello, 6. E-36208 Vigo, Spain
2
ABSTRACT
Concentration and fractionation patterns of rare earth elements (REE) have been studied in sediments from the
Tagus estuary. The spatial distribution pattern of REE and PAAS-normalized ratios shows two distinct areas: 1)
the upper estuary and, 2) the middle and lower estuary. The upper estuary is marked by coarse-grain sediment,
with lower ΣREE concentrations. The PAAS-normalized ratios profile shows slight Light REE (LREE)enrichment and a positive Eu anomaly. The middle and lower estuarine sediments are mainly composed by silt
and clay with higher ΣREE concentrations and a clear Medium-REE (MREE)-enrichment relative to LREE and
Heavy-REE (HREE). Concentrations of REE in Tagus were mainly ruled by sediment nature and anthropogenic
sources.
INTRODUCTIÓN
The rare earth elements (REE) have been intensively
studied as natural tracers of biogeochemical processes [1].
Due to their consistent behaviour the REE are widely used
as tracers of sources and processes controlling trace
element distribution in marine sediments [2, 3, 4, 5]. The
REE distribution in sediments is largely controlled by
scavenging processes [6, 7, 8, 9], by redox conditions of
the overlying water column [10], by composition of the
terrigenous source [11] and by potential anthropogenic
inputs [12]. Different works reported on anomalous REE
concentrations in river and marine sediments [12, 13, 14,
15, 16] caused by unnatural liquid or solid inputs derived
from human activities (e.g. industrial plants, acid mine
drainage, agricultural activities). The Tagus Estuary, one of
the largest in Europe (320 km2 total area) has been
contaminated mainly by two industrial areas located in the
north and south margins [17, 18], and domestic effluents
from the metropolitan area of Lisbon [19]. High levels of
metals (As, Pb, Zn, Cu and Hg) have been reported in
surface sediments and sediment cores [17, 18, 20], in
suspended particulate matter near the sources and of the
lower estuary [17, 21, 22]. The aim of this work was to
study the REE distribution and frationating in surface
sediments from the Tagus estuary and to assess rule of
urban and industrial activities as sources of anomalous
REE concentrations.
MATERIAL AND METHODS
Surface sediments (0-5 cm layer) were sampled in the
Tagus estuary using a Van Veen grab sampler and dried at
40° C, sieved through a 2-mm mesh and grounded with an
agate mortar. Samples were completely mineralized and
analysed by inductively-coupled plasma mass spectrometry
(ICP-MS) using a Thermo Elemental, X-Series, following
[23]. The precision and accuracy of the analytical
procedures was controlled through repeated analysis of the
studied elements in certified reference materials: AGV-1
and MAG-1 (USGS).
RESULTS AND DISCUSSION
Total REE concentrations (ΣREE) ranged between 30 and
145 mg.kg-1. The lowest ΣREE concentrations were found
in the coarse-grained sediments sampled in T8, T9 and T10
(upper estuary), with values of 30, 35 and 39 mg.kg-1,
respectively. Conversely, middle and lower estuary
sediments, mainly composed of silt and clay, showed
higher ΣREE concentrations, varying from 103 to 145
mg.kg-1. These results suggest the relevance of sediment
particle nature in the transport of REE, assuming that
contents in terrigenous sediments increase in the sand-siltclay series [24]. Light REE (LREE) concentrations are
higher than Heavy REE (HREE) for all surface sediments
studied in the estuary.
The REE concentrations were normalized to post-Arcaen
Australian Shales (PAAS) [25] to allow one to identify
within the typical sedimentary REE patterns, an enrichment
or deficiency of a single element or group of elements [26].
PAAS-normalized ratios show two distinct REE
fractionation patterns. The first pattern was observed in the
upper estuary (T8, T9 and T10), with a slight enrichment of
LREE relative to HREE. The (La/Yb)PAAS values vary from
1.32 to 1.61, as expected in normal estuaries with mixing
28
(2016)
between neutral and/or slightly basic waters, where REE
through salinity-induced precipitation are removed in the
following order LREE>MREE>HREE [27]. A positive
Europium anomaly ([Eu/Eu*]) ranging from 1.57 to 2.10,
indicate a significant Eu anomaly in these sediments. Such
high anomaly is not common in sedimentary environments
suggesting changes in the redox cycle of this element
and/or anthropogenic input [28]. A second REE
fractionation pattern was observed in sediments from the
middle and lower estuary (T1-T7), revealing clearly a
MREE-enrichment profile. The (La/Gd)PAAS values,
varying from 0.68 to 0.73, suggest a different input of
particles with a distinct origin. In this section of the estuary
has been previously reported an hotspot for many
contaminants [29] related to the existence of old chemical
and metallurgic activities, namely those using mineral
pyrite ores as raw material.
1,20
T1
T2
T3
T4
T5
T6
T7
T8
T9
T10
1,00
REE/PAAS
0,80
0,60
0,40
0,20
0,00
La Ce Pr Nd Sm Eu Gd Tb Dy Ho Er Tm Yb Lu
Fig. 1. PAAS-Normalized REE patterns for surface
sediments collected in the Tagus estuary.
ACKNOWLEDGEMENTS
The current work is part of the FCT project PTDC/QEQEPR/1249/2014.
REFERENCES
1 - Oliveri, E., Neri, R., Bellanca, A., Riding, R., 2010.
Sedimentology, 57: 142–161.
2 - Piper, D.Z., 1974. Chem. Geol., 14: 285–304.
3 - Sholkovitz, E.R., 1990. Chem. Geol., 88: 333–347.
4 - Murray, R.W., Buchholtzten Brink, M.R., Gerlach,
D.C., Russ, G.P., Jones, D.L., 1991. Geochim. Cosmochim.
Acta, 55: 1875–1895.
5 - Censi, P., Incarbona, A., Oliveri, E., Bonomo, S.,
Tranchida,
G.,
2010.
Palaeogeography,
Palaeoclimatology, Palaeoecology, 292, (1-2): 201–210.
6 - Elderfield, H., Greaves, M.J., 1982. Nature, 296: 214–
219.
7 - Whitfield, M., Turner, D.R., 1987. In: Stumm, W. (Ed.),
Aquatic Surface Chemistry. Wiley, New York, pp. 457–
493.
8 - Elderfield, H., 1988. Philos. Trans. Roy. Soc. Lond. Ser.
A, 325: 105–126.
9 - Haley, B.A., Klinkhammer, G.P., McManus, J., 2004.
Geochim. Cosmochim. Acta, 68: 1265–1279.
10 - Liu, Y.-G., Miah, M.R.U., Schmitt, R.A., 1988.
Geochim. Cosmochim. Acta, 52: 1361–1371.
11 - Taylor, S.R., McLennan, S.M., 1985. Blackwell
Scientific Publications, Oxford.
12 - Olmez, I., Sholkovitz, E.R., Hermann, D., Eganhouse,
R.P., 1991. Environ. Sci. Technol., 25: 310–316.
13 - Ravichandran, M., 1996. Mar. Poll. Bull., 32: 719–
726.
14 - Borrego, J., López-González, N., Carro, B., LozanoSoria, O., 2004. Mar. Poll. Bull., 49: 1045–1053.
15 - Oliveira, S.M.B., Silva, P.S.C., Mazzilli, B.P., Favaro,
D.I.T., Saueia, C.H., 2007. Appl. Geochem., 22: 837–850.
16 - Oliveira, M.L.S., Ward, C.R., Izquierdo, M., Sampaio,
C.H., de Brum, I.A.S., Kautzmann, R.M., Sabedot, S.,
Querol, X., Silva, L.F.O., 2012. Sci. Total Environ., 430:
34–47.
17 - Vale, C., 1990. Sci. Total Environ., 97 (98): 137–154.
18 - Canário, J., Vale, C., Caetano, M., 2005. Marine
Pollution Bulletin, 50: 1142–1145.
19 - Canário, J., Vale, C., 2007. Scientific Report,
IPIMAR, June 2007, p. 78.
20 - Caçador, I., Vale, C., Catarino, F., 1996. Estuarine
Coastal Shelf Science, 42 (3): 393–403.
21 - Vale, C., Ferreira, A., Micaelo, A., Caetano, M.,
Pereira, E., Madureira, M., Ramalhosa, E., 1998. Water
Science Technology, 37: 25–31.
22 - Canário, J., Vale, C., Nogueira, M., 2008. Applied
Geochemistry, 23: 519–528.
23 - Prego, R., Caetano, M., Bernardez, P., Brito, P.,
Ospina-Alvarez, N., & Vale, C., 2012. Continental Shelf
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24 - Sholkovitz, E.R., 1988. Am. J. Sci., 288: 236–281.
25 - McLennan, S.M., 1989. In: Lipin, B.R., McKay, G.A.
(Eds.), Geochemistry and Mineralogy of Rare Earth
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26 - Henderson, P., 1984. Elsevier, New York, pp. 1–32.
27 - Sholkovitz, E.R., Szymczak, R., 2000. Earth and
Planetary Science Letters, 179: 299–309.
28 - Sverjensky, D. A, 1984. Earth Plant. Science Lett., 67:
70-78.
29 - Vale, C., J. Canário, M. Caetano, J. Lavrado e P. Brito,
2008. Mar. Pol. Bull., 56:1364-1367.
29
(2016)
Historical Record of Trace Elements (1983-2007) in Scales from Atlantic
Salmon (Salmo salar): Study of Past Metal Contamination from a Copper
Mine (Ulla River and Estuary, NW Iberian Peninsula)
Antonio Cobelo-García1, Paloma Morán2, Clara Almécija1 & Pablo Caballero3
1
Instituto de Investigacións Mariñas de Vigo (IIM-CSIC)
Universidade de Vigo
3
Servizo de Conservación da Natureza de Pontevedra, Xunta de Galicia
2
ABSTRACT
The Ulla river and its estuary was impacted by the Cu mine of Touro which was in operation from 1973 until it
closure in 1988 due to decreasing quality of the mineral and drop of the copper prices. In the present work, we
studied the historical metal contamination in this watershed (1983-2007) by means of the analysis of trace metals
in scales of salmon (Salmo salar). This approach has been widely used for the reconstruction of the environmental
conditions, since scales are permanent records of the influence of exogenous factors.
Results indicate the presence of a significant contamination for several metals (here are given the examples for
Cu, Zn, Au and Ag) during the 1980’s. Concentrations of Cu in salmon scales during the operation of the mine
were 7-15 higher than current values. Since the metal concentration in the scale is proportional to the
concentration in water, we may estimate that concentrations up to 75 nM were typical during the 1980’s. Such
concentrations have been shown to produce adverse sensory and behavioral effects to salmonids. Also important
was the contamination recorded for Au, with concentrations 15-fold higher than in the recent times. This metal,
which is not normally included in environmental monitoring programs, should be taken into account in future
studies.
INTRODUCTIÓN
Mining is one of the main causes of environmental
pollution by heavy metals that pose serious risks to many
aquatic organisms causing mortalities and disrupting
behavior [1]. Touro mine was an open pit in Galicia,
Northwest Spain located in the River Brandelos, tributary
of the Ulla (8° 20′ 12.06″ W 42° 52′ 46.18″ N). Copper was
extracted from 1973 and 1988. This activity led to a
profound degradation of the environment [2]. The copper
extracted at this mine was in the form of pyrite, pyrrhotine
and calcopyrite included in granitic amphibolite. The rock
is rich in iron (Fe) and copper (Cu) sulphides. The heavy
metals existing in the rock (principally Cu and Fe) increase
its solubility at low pH, and the dissolved metals thus pass
into the drainage water. The run-off water was very acidic
and contained high concentrations of sulphates, chlorides
and heavy metals in solution [3]. In 1988 the mine was
closed due to the decreasing quality of the mineral and
falling copper prices throughout the world. Restoration
measures started at the beginning of 2003 with the addition
of tech-nosols, soils of technical origin constituted by
human-made material.
Atlantic salmon (Salmo salar) is a migratory specie
inhabiting river Ulla. They grow in the river for a period of
one-two years after that they undergo smoltification and
emigrate from the river to the sea. Salmon scales have been
used for the reconstruction of the environmental conditions
of their habitat, as they are permanent records of the
influence of endogenous and exogenous factors on their
apatite–protein structure [4].
The aim of this study is, therefore, to reconstruct the
historical metal contamination in the River Ulla (19832007) and its estuary due to the exploitation of the Touro
mine, by means of the metal analysis of salmon scales.
MATERIAL AND METHODS
Stored, dry scales with non-regenerated nucleus were
selected for analysis. Scales were digested with
concentrated ultra-pure HNO3 (Merck Suprapur). Before
digestion, scales were intensively scrubbed and soaked in
MQ water in order to remove surface contamination.
Several scales amounting up to 50-100 mg were selected
for each digestion. Analysis of trace metals in the sample
digests were carried out by means of ICPMS (X Series,
Thermo Elemental), whereas for major elements ICP-OES
(Perkin Elmer Optima 4300DV) was used.
RESULTS AND DISCUSSION
30
(2016)
These findings indicate the existence of a previous
significant contamination in the River Ulla and its estuary
by several trace metals due to the presence of a Cu mine. A
severe contamination was observed for a metal (Au) which
is not normally included in current environmental
monitoring studies and should be taken into account in
future studies.
50
40
Au (ng g-1)
Fish scales consist of a distal layer composed of an organic
framework
impregnated
with
hydroxyapatite,
Ca5(PO4)3(OH), and a proximal layer that is an uncalcified
fibrillary plate [5]. In this study, Ca (9.4 ± 1.0 %, w/w) and
P (6.6 ± 0.7 %, w/w) analysis of the salmon scales indicate
that, roughly, only about 25% of the scale is composed of
hydroxyapatite being the rest organic material.
The influence of the Cu mine during its operation on metal
concentrations in the salmon scale is clearly observed in
the temporal trend of this metal; accordingly, the Cu
concentrations in the 1980’s (15-30 µg/g) exceed about 715 times the values found in the recent years (~ 2 µg/g).
Given that metal concentrations in scales are proportional
to the water concentrations [6], and taking a ‘pristine’
value of dissolved Cu of 1-5 nM in the present days, we
may estimate copper concentrations up to 75 nM (4.8 ppb)
for this river in the 1980’s. Such concentrations have been
shown to produce adverse sensory and behavioral effects to
salmonids [7].
Importantly, this contamination from the Cu mine was also
significant for other metals, such as Zn (Figure 1), Au and
Ag (Figure 2).
30
20
10
0
1980
1985
1990
1995
2000
2005
2010
2000
2005
2010
Year
30
25
Ag (ng g-1)
30
Cu ( g g-1)
25
20
20
15
10
15
5
10
0
1980
5
1985
1990
1995
Year
0
1980
1985
1990
1995
2000
2005
2010
Year
160
ACKNOWLEDGEMENTS
The technical assistance of S. Calvo (IIM-CSIC) during the
analysis of metals in scales is greatly acknowledged.
140
Zn ( g g-1)
Fig. 2. Temporal variation of Au and Ag concentrations in
salmon scale from the Ulla River
120
REFERENCES
100
1 Scott, G. R. & Sloman, K. A. (2004). Aquatic toxicology,
68, 369-392.
2 Otero, X. L., et al (2012). Journal of Geochemical
Exploration, 112, 84-92.
3 Fernandez, J. A., et al (2006). Environmental Pollution,
139, 21-31.
4 Farrell, A.P., et al (2000). Archives of Environmental
Contamination and Toxicology, 39, 515-522.
5 Flem, B., et al (2005). Applied Spectroscopy, 59, 245251.
6 Sauer, G.R., Watabe, N. (1984). Aquatic Toxicology, 5,
51-66.
7 Hecht, S.A., et al (2007). NOAA Technical
Memorandum NMFS-NWFSC-83
80
60
1980
1985
1990
1995
2000
2005
2010
Year
Fig. 1. Temporal variation of Cu and Zn concentrations in
salmon scale from the Ulla River
For example, the average Au values for the period 19831990 was 31 ± 12 ng/g, whereas for the more recent years
(1995-2007) it was 2.1 ± 1.2 ng/g, indicating a 15-fold
increase during the 1980’s with respect to current values.
31
(2016)
Behaviour of Arade Estuary, south of Portugal during summer conditions longitudinal, vertical and horizontal patterns
Cátia Correia1, Alexandra Cravo1 & José Jacob1
1
CIMA, FCT, Universidade do Algarve, Campus de Gambelas, 8005-139 Faro. [email protected]
ABSTRACT
Arade River is the second most important river in the Algarve coast, depicting a high anthropogenic pressure
mainly during the summer season. This work aims at a better understanding the behaviour of the (lower) Arade
Estuary during summer conditions regarding its physical-chemical characteristics (temperature, salinity, pH,
dissolved oxygen saturation, nutrients, chlorophyll a and suspended solids). A longitudinal transect from the
mouth to 7 km upstream was carried out during a spring tidal cycle together with the sampling of vertical profiles
taken along a cross-section of the lower estuary. Data analysis reveals that only nitrate and silicate have a
conservative behaviour along the estuary, while phosphate and ammonium showed departures from the mixing
line particularly close to a station near the sewage treatment plant in the Portimão margin. Moreover, these data
reveal that the river is contaminated by nitrate since there, high concentration (~200 µM) were reached. Across the
section, vertically there was no stratification of the waters while horizontally the Portimão margin show to have
the highest nutrients and suspended solids concentrations, particularly during the ebb to low tide. During this
campaign chlorophyll a was relatively high, reaching ~2 µg L-1. The evaluation of mass exchanges trough the
cross section reveals that the Arade Estuary may supply chlorophyll a, suspended solids and nutrients to the coast,
contributing to increase its biological productivity.
Estuaries are areas of high scientific, ecologic and
economic values. These systems are recognized as
productive systems, providing important habitats for local
wild life and protection for a large number of species.
However, these are also highly affected by anthropogenic
pressure. Estuaries are transition zones where freshwater
meets salty, sea waters, exhibiting strong physical and
chemical gradients [1]. The quantity of fresh water that
enters the estuary through the river influences the ecology
of the system, promoting gradients and fluctuations
throughout its extension, particularly at the level of
salinity, temperature, oxygen and nutrients [2].
Monitoring these systems is essential for determining its
environmental quality and understanding how these
interplay with the adjacent coastal areas [3]. Arade estuary
[4] is one of the most important coastal systems in the
south of Portugal, where the chemical variability is poorly
studied, and programs of water quality monitoring are
scarce. The Arade river has ~ 75 km extension and has a
drainage basin of 966 km2 [5]. It shows a mesotidal
regime, where the salt water intrusion comes up 16 km
upstream of the mouth, with a mean depth of 6 m and a
maximum depth of 10 m in the mouth, near the city of
Portimão [6].
lower estuary, from Portimão to Parchal margins (250 m),
along a tidal cycle, to understand and characterize the
vertical and horizontal fields of nutrients. The chemical
parameters selected, temperature, salinity, pH and O2
(dissolved and saturation percentage), were measured in
situ with a YSI 6820 multi-parametric probe) and water
samples were collected for suspended solids, chlorophyll
a and nutrients (ammonium, nitrate, phosphate and
silicate) analyses. The water column depth of the section
varied from < 1.5 m at low tide to 6 m at high tide,
greatest at the central area. At the laboratory, the samples
for the determination of suspended solids were filtered
through a membrane filter (0.45 µm) previously
decontaminated and weighted, and then dried to 100ºC 105ºC for an hour. The filtered water was frozen to -20 ºC
for posterior analysis of nutrients. Samples for chlorophyll
a concentration were filtered through GF/F glass fiber
filters (0.7 µm), frozen to -20ºC before analysis. For the
determination of nutrients and chlorophyll a the
spectrophotometric methods described by [7] and [8] were
applied, respectively, and for the SS determination the
gravimetric method described by [9] was applied.
Statistical Analysis
Significant differences at a level of 95% confidence were
sought vertically and horizontally throughout the section
and along the time between samplings.
MATERIAL AND METHODS
RESULTS AND DISCUSSIÓN
To understand the behavior of the nutrients, 8 stations
along the lower estuary upstream (7 km) were
characterized during the ebb period of a spring tide.
Additionally, a crosssection was also sampled on the
The longitudinal transect shows that upstream the
chlorophyll a concentration is relatively high (~ 2 µg L-1)
for a late summer situation. The same was observed at the
section in the afternoon period. The theoretical dilution
INTRODUCTION
32
(2016)
line (TDL) for the nutrients shows that only the nitrates
and the silicates have a conservative behavior (Fig. 1),
while the phosphates and ammonium display a similar
behavior throughout all stations, with departures from the
TDL close to the station near the wastewater treatment
plant. The application of the TDL also allows the
estimation of the concentration of nitrates and silicates
where S = 0, showing that while the silicates
concentration is typical (~150 µM), for nitrates there is a
contamination, as these reached values up to ~200 µM.
This fact is due to the anthropogenic pressure that is felt,
either by urban or agricultural influence.
a
b
; r = 0.99
d
c
; r = 0.95
Fig. 1. Variation of (a) Ammonium, (b) Nitrate + Nitrite,
(c) Phosphate and (d) Silicate concentrations in the 8
stations characterized during the ebb period.
Across the section, data show that there is no stratification
in the water column (p > 0.05) and, therefore, only the
mean is represented for the several variables (Fig. 2).
However, in this section, significant temporal fluctuations
occurred along the tidal cycle (p < 0.05) for the several
parameters measured, as can be seen in Fig. 2, for the
extreme conditions of high water (HW) and low water
(LW). For the nutrients, an antiphase pattern of variation
with the tide due to the dilution effect was observed,
occurring during the flood period, when the
concentrations were minimum (Fig. 2).
a
b
Fig. 2. Mean variation of Ammonium, Nitrate + Nitrite,
Phosphate and Silicate concentrations along three sites
selected in the section during: (a) high water and (b) low
water.
In situ, temperature ranged between 18 ºC (LW) and 22 ºC
(HW), and salinity varied between ~30 (LW) and 36
(HW). The oxygen was under-saturated (< 100%) at LW,
particularly at the Portimão margin, while at HW it was
supersaturated over the entire section (100-105%, not
shown). Horizontally, although the gradients were not
strikingly evident, it is possible to observe significant
differences (p < 0.05) between the two stations closer to
the Portimão margin (50-100 m) in relation to the farthest
station, in the other margin (250 m), like for temperature,
salinity and % of saturation (not shown). For the
chlorophyll a, suspended solids and nutrients, despite no
significant differences (p > 0.05) occurred, values were
higher in the Portimão margin (Fig. 2), associated to a
stronger anthropogenic influence along the city of
Portimão (as the wastewater treatment plant) leading to
these increases. Furthermore, the horizontal heterogeneity
of data along the section could be associated to the flow
patterns of water, suggesting differences in circulation
between the two margins. If it is assumed that this section
of the lower estuary has an ebb behaviour, it can be
supposed that in this period of the year there was an
export of chlorophyll a, suspended solids and nutrients,
contributing to the increase of the biological productivity
of the adjacent coastal zone.
ACKNOWLEGEMENTS
The authors would like to thank to APSines –
Administração dos Portos de Sines e do Algarve S. A., in
particularly to Engª Filipa Duarte for the logistic support
provided through the campaign and to all the team that
participated in the campaigns.
REFERENCES
1 – Statham, P. J., 2012. Nutrients in estuaries – An
overview and the potential impacts of climate change. Sci.
Total Environ. 434, 213-227.
2 - Fujii, T., 2007. Spatial patterns of benthic macrofauna
in relation to environmental variables in an intertidal
habitat in the Humber estuary, UK: Developing a tool for
estuarine shoreline management. Estuar., Coast. Shelf.
Sci. 75, 101-119.
3 – Kramer, K. J. M., 1994. Biomonitoring of coastal
waters and estuaries. CRC Press.
4 - Gomes, A. I., 2013. Alterações Ambientais na Costa
Algarvia durante o Holocénico: Um Estudo com base em
diatomáceas. Universidade do Algarve.
5 – SNIRH, 2013. Sistema Nacional de Informação de
Recursos Hídricos (1995 - 2013). Online: http://snirh.pt/
6 – MARETEC, 2014. Definição do limite jusante dos
estuários portugueses.
Online:http://www.maretec.mohid.com/Estuarios/MenusE
stuarios/Arade_Menu.htm
7 – Grasshof, K., Erkhardt, M., Kremling, K., 1983.
Methods of Seawater Analysis. Verlag Chemie, New
York.
8 – Lorenzen, C., 1967. Determination of chlorophyll and
pheopigments: spectrophotometric equation. Limnol.
Oceanogr. 12, 343 – 346.
33
(2016)
9 – American Public Health Association, American Water
Works Association Water Environment Federation, 1992.
Standard Methods for the Examination of Water and
Wastewater.
Maryland,
USA.
34
(2016)
Plástico en mares y océanos: un problema global solucionable.
Andrés Cózar1
1
Departamento de Biología, Facultad de Cc. del Mar y Ambientales, Universidad de Cádiz, Campus de Excelencia
Internacional del Mar (CEIMAR), E-11510 Puerto Real, Spain
RESUMEN
Resulta muy significativo que, con solo unas décadas de uso generalizado de materiales plásticos, el hombre haya
inundado con residuos plásticos todos los océanos. La contaminación marina por plásticos es uno de los asuntos
que mejor ilustra la capacidad del hombre para modificar la apariencia y composición del planeta. La enormidad
de los océanos parecía suficiente para diluir nuestros desechos, pero nos hemos encontrado en poco tiempo con un
problema de escala planetaria. El progresivo incremento en la producción global de plástico y nuestra dependencia
de este material hacen además pensar que se trata de un problema de difícil solución. Por otra parte, la
contaminación marina por plásticos es un asunto que ha conseguido conectar ciencia, medios de comunicación y
sociedad como en pocas ocasiones. En las últimas dos décadas, esta comunión se ha reforzado y retroalimentado
de tal forma que el número de trabajos de investigación así como el grado de implicación social ha crecido
exponencialmente, un movimiento que empieza incluso a influenciar las estrategias del propio sector empresarial
del plástico. La contaminación por plástico puede por tanto llegar a convertirse también en un ejemplo de la
capacidad del hombre para afrontar y corregir los problemas ambientales que amenazan nuestro planeta. En esta
charla, mostraremos una panorámica integral de la problemática de la contaminación marina por plástico,
mostrando los avances más recientes y las perspectivas de futuro para su estudio y gestión.
INTRODUCCIÓN
La acumulación de residuos plásticos en mares y océanos
es un problema que genera gran preocupación social
debido a los numerosos ejemplos de impactos sobre
organismos así como las evidencias científicas que
demuestran la escala planetaria de esta contaminación. Se
han documentado impactos por ingestión y enredamiento
en invertebrados, peces, aves, tortugas, y hasta grandes
cetáceos [1,2]. Se han encontrado acumulaciones de
residuos plásticos en costas, fondos y aguas de casi todas
las regiones del planeta [3, 4]. Existe además gran
incertidumbre acerca de los posibles efectos de la
contaminación marina por plásticos a nivel ecosistémico
[4, 5] o incluso en la salud humana [1].
El escenario descrito unido al incremento exponencial en
la producción global de plástico hacen pensar que los
esfuerzos por combatir esta contaminación han sido
infructuosos, y que este es un problema de difícil solución
que puede deparar consecuencias a gran escala [6].
RESULTADOS Y DISCUSIÓN
Es sabido que los desechos plásticos marinos se pueden
acumular en aguas superficiales e intermedias, costas,
fondos, e incluso en la biota. Sin embargo, el único stock
de plástico que ha podido ser evaluado a escala global es
el la capa superficial del océano, gracias al uso extensivo
de redes de arrastre superficial para medir concentraciones
de plástico flotante. Análisis de amplia escala espacial y
temporal para otros reservorios de plástico son todavía
difíciles de abordar.
Año tras año se completa el mapa global de residuos
plásticos flotantes (Fig. 1). Las zonas de convergencia de
cada una de las cinco Giros Subtropicales se han
identificado como grandes regiones de acumulación de
desechos flotantes. Los modelos de circulación oceánica
predicen potenciales de acumulación de plásticos en mares
semi-cerrados con alta densidad poblacional, lo que ha
sido demostrado recientemente para el caso del Mar
Mediterráneo [7]. La posibilidad de acumulación de
plástico en las latitudes polares ha sido hasta ahora
pasada por alto, aunque una reciente expedición
circumpolar ha permitido completar esta parte del mapa
global con resultados sorprendentes.
En la dimensión temporal, las series históricas de
contaminación por plásticos flotantes, disponibles desde
los años 80 para algunas regiones [8, 9], convergen en la
conclusión de que no aparecen claras tendencias de
incremento en el grado de contaminación durante los
últimos años, un resultado que no ha sido explorado en
profundidad por la comunidad científica. De hecho,
existen grandes incógnitas en relación a cuál ha sido
35
(2016)
realmente la evolución histórica de la contaminación
marina por plástico y cómo las medidas aplicadas para su
gestión han incidido en las tendencias históricas.
Los desechos plásticos sufren un continuo proceso de
fragmentación que hace que se puedan encontrar en el mar
desde objetos del orden de metros a partículas de pocas
micras. Debido a esta movilidad en la escala de tamaños,
las afecciones sobre organismos y ecosistemas pueden
llegar a ser muy diversas, actuando a múltiples niveles [12, 5]. Especialmente llamativos son los nano-plásticos (en
la escala de micras), capaces de ser incorporados en el
tejido de sus consumidores [10], lo que plantea una vía
potencial de impacto totalmente desconocida.
Estos y otros avances en el conocimiento de la
contaminación marina por plásticos han atraído
enormemente la atención de medios y ciudadanía a nivel
global. Así, numerosas iniciativas sociales se han puesto
en marcha para combatir la contaminación marina en los
últimos años. La sinergia entre el conocimiento científico
y la implicación social crece, existiendo ejemplos de
actuaciones preventivas y correctoras a escala regional de
enorme éxito. El fenómeno de la contaminación por
plástico es un problema inquietante, pero la preocupación
y la acción social surgida hacen pensar que esta amenaza
global pueda ser solucionable.
AGRADECIMIENTOS
Esta contribución es el resultado de la colaboración de una
larga lista de investigadores en distintos proyectos
(Malaspina CSD2008-00077, MedSeA FP7-2010-265103;
Programa S. de Madariaga PRX14/00743, Tara Oceans)
Fig. 1. Primer mapa de contaminación marina por
plásticos flotantes (Fuente: National Geographic, Autores:
A. Cózar y J. Hawk).
REFERENCIAS
1 - Rochman CM, et al. 2015. Anthropogenic debris in
seafood: Plastic debris and fibers from textiles in fish and
bivalves sold for human consumption. Sci. Rep., 5:14340.
2 - de Stephanis R, et al. 2013. As main meal for sperm
whales: Plastics debris. Mar Pollut Bull, 69(1–2):206–
214.
3 - Pham CK, et al. 2014. Marine litter distribution and
density in European seas, from the shelves to deep basins.
PLOS ONE 9(4): e95839.
4 - Cózar A, et al. 2014. Plastic debris in the open ocean.
PNAS 2014 111(28): 10239–10244.
5 - Sussarellu R., et al. 2016. Oyster reproduction is
affected by exposure to polystyrene microplastics. PNAS
113(9): 2430-2435.
6 - Wilcox, C., et al. 2015. Threat of plastic pollution to
seabirds is global, pervasive, and increasing. PNAS
112:11899–11904.
7 - Cózar A, et al. 2015. Plastic accumulation in the
Mediterranean Sea. PLOS ONE 10(4): e0121762.
8 - Law KL et al. (2010) Plastic accumulation in the North
Atlantic Subtropical Gyre. Science 329:1185–1188.
9 - Law KL, et al. (2014) Distribution of surface plastic
debris in the eastern pacific ocean from an 11-year data
set. Environ Sci Technol 48(9):4732–4738.
36
(2016)
10 - Avio CM, et al. 2015. Pollutants bioavailability and
toxicological risk from microplastics to marine mussels.
Environ Pollut., 198:211-22.
37
(2016)
Simulating CO2 leakage from sub-seabed storage to determine metal toxicity
in marine bacteria
Alejandra Díaz1, Ana R. Borrero-Santiago1*, T. Ángel DelValls1 & Inmaculada Riba1
1
Departamento de Química-Física, Facultad de Ciencias Del Mar y Ambientales, Universidad de Cádiz, UNESCO/UNITWIN
Wicop, Polígono Río San Pedro s/n, 11510 Puerto Real, Cádiz, Spain
RESUMEN
La captura y almacenamiento de CO2 (CAC) en formaciones geológicas estables es considerada una de las
estrategias más adecuadas para disminuir las emisiones directas de CO2 en la atmósfera. Sin embargo, esta técnica
alberga cierta incertidumbre ante un posible riesgo de fuga de CO2. Estudios previos han obtenido resultados
adversos en organismos marinos, observándose un aumento de estos efectos cuando se trata de sedimento
contaminado por metales, debido a la movilización durante el proceso de acidificación. En este aspecto, las
comunidades bacterianas del sedimento han sido poco consideradas. Por tanto, este estudio simuló posibles fugas
de CO2 utilizando un sistema de inyección de CO2 asociado a una cámara adaptada para trabajar con
microrganismos; y con ello evaluar la toxicidad asociada a la movilidad de metales como el Zn y el Cd en dos
poblaciones distintas de bacterias (Roseobacter sp. y Pseudomonas litoralis). Las respuestas de las dos
poblaciones se determinaron a partir de variables como el número total de células (células·mL-1), ratios de
crecimiento (μ, hora-1), efecto inhibitorio del CO2 (RICO2), y producción de sustancias exopoliméricas (EPS) (μg
Glucosa·células-1). De manera general se observó un efecto negativo en todas las variables estudiadas a medida
que el pH disminuía. Además se obtuvieron diferencias significativas entre las dos poblaciones en función del
metal (Zn y Cd).
INTRODUCCIÓN
La urgente necesidad de reducir la emisión de CO2 ha
llevado a la búsqueda de soluciones innovadoras. Una de
las medidas consideradas para este fin es la captura y
almacenamiento de CO2 (CAC) en formaciones
geológicamente estables1. Este tipo de tecnología puede
llevarse a cabo en zona terrestre o marítima2. Sin embargo,
existen riesgos potenciales asociados a esta actividad3. En
el caso de un posible escape de CO2 procedentes de una
zona de almacenaje marina, el paso del gas provocaría un
cambio en el equilibrio físico-químico en el agua
intersticial del sedimento debido a la acidificación. Estos
cambios incluyen la alteración de parámetros como el pH,
así como en la movilidad, especiación y, por tanto,
biodisponibilidad de metales del sedimento marino al
tratarse de sedimentos contaminados por metales4. En este
contexto, la combinación de los efectos ante una
acidificación y la biodisponibilidad de metales derivados
de CAC en poblaciones bacterianas marinas aún no han
sido estudiados5. Estudios previos reconocen la tolerancia o
las adaptaciones de las bacterias marinas a lugares
contaminados. No obstante, hasta ahora no se ha descrito sí
esta capacidad podría verse afectada ante un escape de
CO2. Este hecho requiere especial atención ya que las
comunidades bacterianas del sedimento son las encargadas
de los procesos de degradación de la materia orgánica y
están asociadas a los ciclos del carbono del océano; y por
tanto cambios de pH provocados por una fuga de CO2
podrían comprometer el equilibrio natural. El objetivo
principal fue evaluar las respuestas de dos poblaciones
bacterianas marinas (Roseobacter sp. y Pseudomonas
litoralis) expuestas a distintas inyecciones de CO2 en
presencia de concentraciones de Zn y Cd registradas en los
sedimentos de la Bahía de Cádiz. Las respuestas de ambas
poblaciones se evaluaron en función de los resultados
obtenidos en número total de células (células·mL-1) al final
del ensayo de toxicidad, ratios de crecimiento (μ, hora-1),
efecto inhibitorio del CO2 (RICO2), evolución de las curvas
de crecimiento y producción de sustancias exopoliméricas
(EPS) (μg Glucosa·células-1)
MATERIAL Y MÉTODOS
Tanto Roseobacter sp. (CECT 7117) como Pseudomonas
litoralis (CECT 7669) fueron obtenidas en forma de
ampolla liofilizada de la Colección Española de Cultivos
Tipo. El medio de cultivo empleado para llevar a cabo los
ensayos de toxicidad fue una dilución del medio Marine
Broth 2216 (Difco) diluído en agua de mar (1:10)
propuesto por Borrero-Santiago et al. (2016). Las distintas
concentraciones de Zn y Cd (204.1 mg·L-1 para Zn and
0.247 mg·L-1 para Cd)6 se obtuvieron por medio de ZnCl2 y
CdCl2. Los tratamientos de pH correspondieron con pH 7.8
(sin inyección de CO2 como control), 7, 6.5, 6 y 5.5.
Además, un control de pH 7.8 negativo en ausencia de Zn o
38
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Cd; y un control de pH 7.8 positivo en presencia de Zn o
Cd. El número de células totales se determinó a través de
una recta de calibrado, relacionando el contaje directo de
células en microscopio de epifluorescencia mediante la
técnica del flourocromo DAPI, con su correspondiente
densidad óptica a 660 nm (OD660). La estimación de las
curvas de crecimiento se realizó en intervalos de 0, 6, 12,
24, 48 y 72 h. Los ratios de crecimiento (µ) fueron
estimados utilizando la ecuación propuesta por Widdle6 y
el efecto inhibitorio de CO2 (RI) con la fórmula propuesta
por Enfors and Molin7. Los exopolisacáridos (EPS) fueron
cuantificados por el método fenol-sulfúrico8 usando
glucosa como estándar.
RESULTADOS Y DISCUSIÓN
Los resultados obtenidos en este estudio mostraron un
incremento en la toxicidad de Zn y Cd a medida que
decrecía el pH. En general, el crecimiento de Roseobacter
sp. y P.litoralis se vio afectado de manera negativa por la
combinación de CO2 en presencia de Zn o Cd. Este hecho
fue reflejado por incrementos en los tiempos de
aclimatación (fase lag), reducción del número de células,
así como en los ratios de crecimiento. Además, efectos
metabólicos asociados a la producción de EPS mostraron
un incremento en la síntesis de EPS como respuesta a la
combinación tóxica de metal y CO2. Ambas especies
fueron capaces de sintetizar EPS aún viendo inhibido su
crecimiento. Por tanto, estos resultados evidencian que la
combinación de eventos de acidificación en zonas
contaminadas por metales podría afectar a comunidades
bacterianas del sedimento que a priori pueden ser
resistentes a esas concentraciones.
AGRADECIMIENTOS
This work was partially supported by the Spanish Ministry
of Economy and Competitiveness under grants: CTM201128437- C02-02/TECNO and CTM2012-36476-C02-01 and
the international Grant from Bank Santander/UNESCO
Chair UNITWIN/WiCop. First author thanks the Erasmus
Mundus Programme for the master scholarship (20140693/001-001- EMJD). Ana R. Borrero-Santiago thanks
the Spanish Ministery of Science and Education for her
F.P.I. fellowship (BES-2012-054950).
REFERENCIAS
1 – IPCC, 2013. Climate Change 2013: The physical
Science Basis. Contribution of Working Group I to the
Fifth Assessment Report of the Intergovernmental Panel on
Climate Change. Intergov. Panel Clim. Chang. Work. Gr. I
Contrib. to IPCC Fifth Assess. Rep. (AR5) (Cambridge
Univ Press. New York) 1535
2 – Hofmann, M. & Schellnhuber, H.J., 2010. Ocean
acidification: a millennial challenge. Energy Environ. Sci.
3, 1883-1896. Doi:10.1039/C000820F
3 – Damen, K. et al., 2006. Health, safety and
environmental risks of underground CO2 storage –
Overview of mechanisms and current knowledge. Clim.
Change 74, 289-318. Doi:10.1007/s10584-005-0425-9
4 – De Orte, M.R. et al., 2014. Metal mobility and toxicity
to microalgae associated with acidification of sediments:
CO2 and acid comparison. Mar. Environ. Res. 96, 136-144.
Doi:10.1016/j.marenvres.2013.10.003
5 – Borrero-Santiago, A.R. et al., 2016. Carbon Capture
and Storage (CCS): risk assessment focused on marine
bacteria.
Ecotoxicol.
Environ.
Saf.
(2016),
Doi:10.1016/j.ecoenv.2016.040.020 (In press)
6 – Rodríguez-Romero, A. et al., 2013. Several benthic
species can be used interchangeably in integrated sediment
quality assessment. Ecotox. Environ. Safe. 92, 281–288.
Doi: 10.1016/j.ecoenv.2013.02.015
7 – Widdel, F., 2010. Theory and measurement of bacterial
growth. In: Di Dalam Grundpraktikum Mikrobiologie, 1–
11.
8 – Enfors, S. O. & Molin, G., 1981. The influence of
temperature on the growth inhibitory effect of carbon
dioxide on Pseudomonas fragi and Bacillus cereus. Can. J.
Microbiol. 27, 15-19. Doi:10.1139/m81-003
9 – Dubois, M. et al., 1956. Colorimetric method for
determination of sugars and related substances. Anal.
Chem. 28(3), 350–356. Doi: 10.1021/ac60111a017
10 – Nweke, C. O. et al., 2007. Toxicity of zinc to
heterotrophic bacteria from a tropical river sediment. Appl.
Ecol. Environ. Res. 5(1), 123–132.
11 – Nielsen, P.H. & Jahn, A., 1999. Extraction of EPS. In:
J. Wingender et al. (eds), Microbial Extracellular
Polymeric Substances , pp. 49–72. ISBN 978-3-642-601477
39
(2016)
Antioxidant activity of eight microalgae strains
Adela Durá1, Isabel Guerra1, Milagros Rico1, Argimiro Rivero1 & Juan Luis Gómez Pinchetti2
1
Grupo QUIMA- Instituto de Oceanografía y Cambio Global, Universidad de Las Palmas de Gran Canaria, Campus de Tafira,
35017 Las Palmas de Gran Canaria, Canary Islands, Spain
2
Banco Español de Algas, Instituto de Oceanografía y Cambio Global, Universidad de Las Palmas de Gran Canaria, Muelle de
Taliarte s/n, 35214 Telde, Canary Islands, Spain
RESUMEN
Estudios recientes han demostrado que las microalgas son una fuente rica en moléculas bioactivas de especial
valor en los ámbitos de la alimentación y la salud humana. Entre los compuestos más destacables presentes en
dichas algas se encuentran los polifenoles, cuya capacidad antioxidante ha suscitado gran interés en los campos
citados. Estos, son capaces de inhibir la acción de los radicales libres, causantes de multitud de enfermedades
crónicas como son el cáncer, enfermedades cardiovasculares y enfermedades neurodegenerativas. Las últimas
investigaciones revelan que algunos antioxidantes sintéticos como el butilhidroxitolueno (BHT) y el
butilhidroxianisol (BHA) empleados como conservantes en las industrias de alimentación y farmacéutica podrían
presentar efectos carcinógenos, lo cual ha potenciado la búsqueda de nuevos y eficientes conservantes naturales.
El objetivo de este estudio consiste en evaluar la capacidad antioxidante que presentan los extractos en metanol,
etanol/agua y agua de ocho cepas de microalgas y cianobacterias diferentes, por medio del método de inhibición
del radical libre 1,1-difenil-2-picrilhidrazil (DPPH). A continuación, se citan los nombres de las especies de
microalgas y cianobacterias seleccionadas y donadas por el Banco Español de Algas (BEA) para esta
investigación: Chloroidium saccharophilum, Pseudopediastrum boryanum, Cosmarium blyttii, Cosmarium sp.,
Pseudopediastrum boryanum, Spyrogyra sp., Ochrosphaera verrucosa y Chloromonas cf. reticulata. Para
identificar y cuantificar los polifenoles responsables de dicha actividad se emplea la técnica de cromatografía
líquida de alta resolución. Finalmente, con el propósito de valorar su posible aplicabilidad comercial como aditivo
alimentario, los resultados obtenidos se comparan con los que ofrecen los antioxidantes sintéticos BHA y BHT.
ABSTRACT
Recent studies have demonstrated that microalgae are a rich source of bioactive molecules of particular value in
the fields of food and human health. One of the most outstanding compounds present in these algae are
polyphenols, which have attracted great interest due to their high antioxidant capacity. Phenolic compounds have
been reported to be great free-radical inhibitors and thus, they are believed to protect against diseases caused by
oxidative stress, such as cancer, cardiovascular diseases and neurodegenerative diseases. Latest data reveal that
synthetically produced antioxidants currently used as preservatives in food and pharmaceutical industries such as
butylated hydroxytoluene (BHT) and butylated hydroxyanisole (BHA) could manifest carcinogenic potential. This
fact has fuelled research on finding new and efficient natural preservatives.
The aim of this study consists in determining antioxidant capacity exhibited by methanol, ethanol/water and water
microalgal extracts of eight different species by using the 1,1-diphenyl-2-picrylhydrazyl (DPPH) method. The
following microalgae strains were selected and donated by the Spanish Bank of Algae (BEA) for their
examination: Chloroidium saccharophilum, Pseudopediastrum boryanum, Cosmarium blyttii, Cosmarium sp.,
Pseudopediastrum boryanum, Spyrogyra sp., Ochrosphaera verrucosa y Chloromonas cf. reticulata. In addition,
the major phenolic constituents present in the extracts were identified and quantified by means of high
performance liquid chromatography. Finally, in order to assess their possible commercial applicability as a food
additive, results of this assay are compared to those provided by synthetic antioxidants BHA and BHT.
AGRADECIMIENTOS
Las microalgas y cianobacterias anteriormente citadas
fueron recolectadas en Canarias y depositadas en la
colección de cultivos del Banco Español de Algas (Taliarte,
España). Los autores de este artículo desean expresar su
gratitud al Banco Español de Algas por la donación de
dichas cepas.
40
(2016)
REFERENCIAS
1 – Rico M, Sánchez I, Trujillo C, Pérez N, 2013.
Screening of the antioxidant properties of crude extracts of
six selected plant species from the Canary Islands (Spain).
J. Appl. Bot. Food Qual. 86:217 – 220.
2 – Rodríguez-García I, Guil-Guerrero JL, 2008.
Evaluation of the antioxidant activity of three microalgal
species for use as dietary supplements and in the
preservation of foods. Food Chem. 108:1023–1026.
41
(2016)
Determination of UV-filters in seawater samples from Gran Canaria island
using fabric phase sorptive extraction (FPSE) coupled to LC-MS/MS
Romualdo Bentor García Guerra, Sarah Montesdeoca Esponda, María Esther Torres Padrón,
Zoraida Sosa Ferrera, José Juan Santana Rodríguez
Grupo de Análisis Químico Medioambiental (AQMA), Instituto Universitario de Estudios Ambientales y Recursos Naturales
(i-UNAT), Universidad de Las Palmas de Gran Canaria, 35017 Las Palmas de Gran Canaria, Spain.
*[email protected], Tel.: +34 928 452 915; Fax: +34 928 452 922.
ABSTRACT
Benzotriazole UV stabilizers (BUVSs) are a group of emerging compounds whose use has increased in the last
decades due to the growing concern about the link between sunlight exposure and skin cancer. After be used,
BUVSs can reach the environment through recreational activities such as swimming and bathing in oceans, lakes
or rivers or after passing throughout wastewater treatment plants without be removed. Fabric phase sorptive
extraction (FPSE) is a new reusable extraction technique, which integrates the advantages of sol-gel hybrid
inorganic-organic sorbents with fabric substrates, being highly sensitive, efficient and cheap device. After
optimize the parameters that affect the procedure (sorbent chemistry, extraction time, sample volume, pH, ionic
strength, back-extraction solvent, time and volume), FPSE coupled to Ultra High Performance Liquid
Chromatography with tandem mass spectrometry (UHPLC-MS/MS) was employed to analyse six BUVSs in
seawater samples from Gran Canaria Island (Spain). The methodology allowed enrichment factors of 25 times
with limits of detection (LODs) from 1.06 to 8.96 pg∙mL-1, recoveries in the range 9.30-51.4% and intra and interday RSDs between 3.97 and 20.8% for all compounds. The application of the proposed procedure to thirty-six
seawater samples from different beaches of the island allowed detecting and quantifying one of the target
compounds in the range from 41.12 to 544.9 pg∙mL-1.
INTRODUCTION
In the last decades, the increase of the human activity and
the rapid industrialization have generated the appearance of
a wide variety of chemical contaminants all over the world.
Among these emerging compounds, benzotriazole UV
stabilisers (BUVSs) added to sunscreen and several
cosmetic such as lip gloss, shampoos, hair dyes, makeup,
etc., have been described as mutagenic, toxic, pseudopersistent and bioaccumulative [1]. BUVSs have a phenolic
group attached to the benzotriazole structure which has a
heterocyclic structure containing three nitrogen atoms
whose molecular form is C6H5N3 [2]. Its entrance to the
environment can be directly through bathing at beaches,
lakes or rivers, or indirectly through discharges from
wastewater treatment plants (WWTPs) [1], where they are
not removed.
Fabric phase sorptive extraction (FPSE), developed by
Kabir and Furton [3], consists on a flexible and permeable
fabric substrate coated with a sol-gel hybrid inorganicorganic sorbent chemically bonded to its surface. It allows
the use of large amounts of sorbent inside the cellulose
substrate, generating a phenomenal increase in the
retention of analytes [4]. FPSE shows advantages such as
simple use and minimal solvent consumption; variety of
sorbents; high pH stability; ease of increasing the diffusion
of the analytes by magnetic stirring or fast analyte.
We have optimized all the parameters that could affect the
extraction efficiency of six BUVSs (UVP, UV360, UV326,
UV327, UV328 and UV329) and then FPSE coupled to
ultra-high performance liquid chromatography and mass
spectrometry detection (UHPLC-MS/MS) was applied to
the analysis of thirty-six samples from different beaches of
Gran Canaria Island (Spain).
MATERIAL AND METHODS
The separation and determination system included an
ACQUITY UHPLC equipped with a binary solvent
manager, a column manager, an autosampler with a syringe
of 25 µL and a tray for 21 HPLC glass vials and a mass
spectrometry detector. An ACQUITY UHPLC BEH C18
column (1.7 µm, 2.1 mm×50 mm) was utilised at 40ºC with
a methanol with 0.1% of formic acid as mobile phase at a
flow rate of 0.5 mL·min−1. Conditions of mass
spectrometry detection were: desolvation temperature,
42
(2016)
450ºC, source temperature, 120ºC, capillary votage, 3kV,
cone votage, 30V and extractor votage, 3V.
The FPSE protocol has been the following: (a) Cleaning of
FPSE media in 2 mL of acetonitrile-methanol (50:50) for 5
minutes, (b) Rinsing with 2 mL of distilled water to remove
residual organic solvents, (c) Extraction with the FPSE
device during the optimum extraction time, stirring the
sample with a magnetic bar at 1000 rpm, (d) Desorption in
the appropriate volume of organic solvent during the
optimum time for complete back-extraction. Then transfer
sample in a deactivated HPLC glass vial for determine the
BUVSs in the chromatographic system. After finishing the
extraction, repetition of the first step, drying the FPSE
media to eliminate the solvents and storage in an air-tight
glass container for future use.
The seawater samples were collected in three different days
and from three beaches at southwest of the island: Mogán,
Amadores and Puerto Rico. Four samples per beach (two
centrals and East and West margins) were collected from
the surface layer (< 10 cm) in an area with a depth of
approximately 1 m.
RESULTS AND DISCUSSION
First, three different FPSE media was tested: Sol-gel poly
dimethyldiphenylsiloxane (PDMDPS), Sol-gel poly
tetrahydrofuran (PTHF) and Sol-gel poly ethylene glycol
(PEG). Sol-gel PDMDPS, which has demonstrated greater
capacity to extract nonpolar compounds like BUVSs [5],
was selected. Second, an experimental design of 24 (four
parameters and two levels) was used to study the influence
of extraction time, sample volume, ionic strength and pH.
Attending to the data obtained, 150 minutes, 25 mL, 5%
(m/v) and pH 6 respectively were selected as optimum
values. Third, two different back-extraction solvents
(methanol and acetonitrile) and a mix of both (50:50) were
tested, the best results being obtained with methanol. Then
we tested the influence of the back-extraction time,
studying values of 5, 10 and 15 min. Highest recoveries of
all the compounds was obtained with 10 minutes. Finally,
three different volumes were used as back-extraction
volume, testing 0.5, 1 and 1.5 mL of methanol. The results
demonstrated that using 1 mL we achieve higher values
and better shape of the chromatographic peaks (Fig. 1).
The validation of the FPSE-LC-MS/MS method was
studied in terms of linearity (correlation coefficients over
0.9932), sensitivity (LODs between 1.06 and 8.96 pg∙mL-1
and LOQs between 3.54 and 29. 9 pg∙mL-1) and intra-day
and inter-day precision (in the range 3.97-10.0% and 5.7120.8%, respectively).
Fig. 1. Chromatograms obtained for each BUVS at 250
ng·mL-1 in the optimised conditions.
The procedure provided enrichment factors of 25 times,
reaching recoveries between 32.4 and 51.4% for all the
studied compounds, except for UV P and UV 329 whose
recoveries were in the range of 9.30 and 21.5%. The
differences between the obtained recoveries may be due to
the different polarities of the target compounds, which
affect their interaction with the FPSE media. Once
optimised the methodology, we determined one of the
studied compounds, UV 360, in nine of the thirty-six
analysed seawater samples from three beaches of Gran
Canaria Island (Spain), in concentrations between 41.12
and 544.9 pg∙mL-1.
REFERENCES
1 - Montesdeoca-Esponda, S., Vega-Morales, T., SosaFerrera, Z. and Santana-Rodríguez, J. J. (2013). Extraction
and determination methodologies for benzotriazole UV
stabilizers in personal-care products in environmental and
biological samples. TrAC Trends in Analytical Chemistry.
51, 23-32.
2 - Herrero, P., Borrull, F., Pocurull, E. and Marcé, R. M.
(2014). An overview of analytical methods and occurrence
of benzotriazoles, benzothiazoles and benzenesulfonamides
in the environment. TrAC Trends in Analytical Chemistry.
62, 46-55.
3 - Kabir, A. and Furton, K. G. (2014). Fabric phase
sorptive extractors (FPSE), US Patent Application:
14,216,121 March 17, 2014.
4 - Racamonde, I., Rodil. R., Quintana, J. B., Sieira, B. J.,
Kabir, A., Furton, K. G. and Cela, R. (2015). Fabric phase
sorptive extraction: A new sorptive microextraction
technique for the determination of non-steroidal antiinflammatory drugs from environmental water samples.
Analytica Chimica Acta. 865, 22-30.
5 - Segro, S. S. and Malik, A. (2008). Solvent-resistant solgel polydimethyldiphenylsiloxane coating for on-line
hyphenation of capillary microextraction with highperformance liquid
chromatography. Journal of
Chromatography A, 1205, 26-35.
43
(2016)
Medida de los flujos bentónicos de materia usando cámaras e incubando
testigos: algoritmos para corregir el artefacto creado por el muestreo del
agua durante su incubación
A.Gómez-Parra, T. Ortega, R. Ponce, J.M. Forja
1
Departamento de Química Física, Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz, Campus Río San
Pedro, 11510 Puerto Real (Cádiz) Spain
RESUMEN
La estimación de los flujos de materia a través de la interfase sedimento-agua usando cámaras bentónicas o
incubando “cores” requiere conocer la variación temporal que experimenta la concentración de la sustancia en
estudio en el agua incubada con el sedimento, lo que implica para la mayoría de los analitos que a lo largo del
tiempo han de extraerse muestras sucesivas para su análisis. Esta operación introduce un importante artefacto ya
que, dependiendo como se opere, o bien el volumen del agua incubada no permanece constante durante todo el
proceso, o su concentración se altera con cada muestreo.
En este trabajo se presentan dos algoritmos desarrollados para la corrección del efecto del muestreo del agua
incubada, tanto si las alícuotas extraídas se sustituyen por agua procedente del exterior, como si su volumen es
ocupado por la expansión de un globo de compensación. En el segundo caso se utiliza un sistema de recurrencia
por ecuaciones en diferencias, lo que permite evitar el error conceptual que supone calcular el flujo a partir de una
regresión lineal de la concentración y el tiempo. Para ambos algoritmos se hace un análisis de los errores
cometidos sin su uso a partir de una amplia base de datos obtenidos con cámaras bentónicas en la plataforma del
golfo de Cádiz y en distintos sistemas litorales de la península Ibérica.
INTRODUCCIÓN
Los flujos de materia desempeñan un papel relevante
dentro del conjunto de procesos biogeoquímicos que tienen
lugar en los sistemas litorales debido fundamentalmente a
su elevada magnitud [1,2]. Entre estos flujos destacan los
provocados por la transferencia de nutrientes y gases del
agua intersticial del sedimento a la columna de agua. En el
primer caso ya que, con frecuencia, no solo superan los
requerimientos nutricionales del fitoplancton en los
ecosistemas intralitorales, sino que exportan nutrientes a
las zonas oceánicas circundantes contribuyendo con ello al
mantenimiento de su productividad primaria [3]. En el
segundo, porque pueden presentar un comportamiento
inverso al tiene el Océano en su conjunto como receptor de
gases con efecto invernadero [4].
La medida directa de los flujos bentónicos se realiza “in
situ” (por medio de cámaras bentónicas) o en el laboratorio
a partir de testigos de sedimento. En ambos casos se
requiere incubar una cierta superficie de sedimento (S,
conocida) en contacto con un determinado volumen aislado
de agua procedente del fondo (V). El cálculo del flujo (J)
se realiza a partir de su propia definición, en la que implíJ=
V dC
·
S dt
citamente V se considera constante.
La expresión se aplica para t = 0 y usando la siguiente
función de la concentración con el tiempo [5]:
– k·t
C = a – (a – C0)·e
Donde a y k son parámetros de ajuste que se obtienen de la
represión de C con t y tienen sentido físico: a es la
concentración de la especie estudiada en el límite superior
del agua intersticial y k es un productorio de constantes de
operación, entre las que se encuentra el volumen de agua
incubado (V) en la cámara o con el testigo de sedimento.
La aplicación de las dos expresiones anteriores puede no
ser posible a consecuencia del muestreo que es necesario
realizar para conocer la dependencia con el tiempo de la
concentración de muchas sustancias que requieren del
análisis químico. La dificultad surge porque la retirada de
las alícuotas que se requieren para los análisis alteran el el
volumen original que se incubaba. Esto podría obviarse
sustituyendo los volúmenes extraídos por otros idénticos de
agua procedente del exterior. No obstante, operando de esta
manera se altera la evolución natural de la concentración
del analito con el tiempo.
44
(2016)
En este sentido, el objetivo de este trabajo es proponer los
algoritmos de cálculo necesarios para corregir el efecto del
artefacto creado por el muestreo del agua incubada con de
las cámaras bentónicas o con los testigos de sedimento,
pudiendo con ello hacer una regresión de C con t
conceptualmente correcta
MATERIAL Y MÉTODOS
El trabajo para la obtención de los flujos que se describe se
realizó por medio de cámaras bentónicas. Se utilizaron tres
tipos de cámaras de diferente tamaño y grado de
automatismo, según las características de las zonas a
estudiar. Así, la superficie de sedimento cubierta varió
entre 0.18 y 0.50 m2 y el volumen de agua incubada entre
47 y 140 L.
misma cuantía que la muestra tomada en cada ocasión, se
considera que la tendencia de la curva C(t) varía cada vez
que se muestra de manera proporcional al volumen
extraído con lo cual se modifica el valor de k. El problema
se resuelve recurriendo un nuevo coeficiente de regresión
k’ (independiente del volumen) cuyo valor, junto con el de
a, permiten obtener el flujo para t = t0, que, al igual que
antes, es justo cuando aún no se ha producido ninguna
alteración por el muestreo.
Los resultados obtenidos muestran que las correcciones a
introducir por el muestreo en la medida de los flujos son
pequeñas. Cuando los volúmenes de agua extraídos son
pequeños los errores cometidos son, en numerosos casos,
del mismo orden de magnitud que las desviaciones de los
replicados medidos cuando se estiman los flujos. No
obstante, a nivel conceptual es importante asumir este
hecho, especialmente en aquellos casos en que sea
necesario sustituir en el agua incubada volúmenes mayores.
AGRADECIMIENTOS
Deseamos expresar nuestro agradecimiento al Dr. J.L. Díaz
Moreno por sus valiosos comentarios acerca de la
expresión de los modelos desarrollados. Este trabajo ha
sido financiado por el proyecto CTM2014-59244-C3.
REFERENCIAS
Fig. 1. Imagen de una de las cámaras
Bentónicas utilizadas.
Las cámaras utilizadas han sido descrita con detalle en
trabajos anteriores [6, 7]. En la figura 1 se muestra una
vista de una de ellas que se utilizó en la plataforma del
golfo de Cádiz (hasta 50 m de profundidad) a bordo del
B/O Mytilus. Las otras, más ligeras, se fondearon desde
embarcaciones de pequeño calado en diversos sistemas
costeros de la bahía de Cádiz, marismas del Palmones
(bahía de Algeciras) y en los estuarios del Tinto-Odiel
(Huelva), Oka (Vizcaya) y Asón (Cantabria).
RESULTADOS Y DISCUSIÓN
Los algoritmos propuestos se refieren a las distintas
maneras en que se pretende solucionar la perturbación que
introduce en el experimento la toma de muestra para
conocer en un instante dado la concentración de la especie
química cuyo flujo se pretende conocer. En el primer caso,
cuando se sustituye el volumen muestreado por agua del
exterior, se produce una dilución del agua incubada. El
problema se aborda calculando la
el valor de la
concentración resultante como la media ponderada de las
masa de agua involucradas admitiendo que la tendencia de
la curva C(t) no varía con cada muestreo. En el segundo
caso, cuando el volumen de la cámara disminuye en la
1 – Jahnke, R., Richards, M., Nelson, J., Robertson, C.,
Rao, A., Jahnke, D., 2005. Organic matter remineralization
and porewater exchange rates in permeable South Atlantic
Bight continental shelf sediments. Cont. Shelf. Res., 25: 1433-1452.
2 – Alongui, D.M., Trott, L.A., Pfitzner, J., 2007.
Deposition, mineralization, and storage of carbon and
nitrogen in sediments of the far northen Great Barrier
Reefshelf. Cont. Shelf. Res., 27:2595-2622.
3 - Gómez-Parra, A., Forja, J.M., 1992. Significance of
benthic regeneration in nutrient balance in the Bay of
Cadiz, south-west Spain (a shallow semi-closed coastal
ecosystem). The Science of the Total Environment,
Supp.1992: 1079-1086.
4 – Borges, A.V., 2005. Do we have enogh pieces of the
jigsaw to integrate CO2 fluxes in the coastal ocean?
Estuaries, 28: 3-27.
5 - Forja, J.M., Gómez-Parra, A., 1998. Measuring nutrient
fluxes across the sediment-water interface using benthic
chambers. Marine Ecology Progress Series. 164: 95-105.
6 - Ferrón, S., Alonso-Pérez, F., Castro, C.G., Ortega, T.,
Pérez, F.F., Ríos, A. F., Gómez-Parra, A., Forja, J.M.,
2008. Hydrodinamic charaterization and performance of an
autonomous benthic chamber for use in coastal systems.
Limnology and Oceanography: Methods, 6: 558-571.
7 - Forja, J.M., Blasco, J., Gómez-Parra, A., 1994. Spatial
and seasonal variation of "in situ" benthic fluxes in the Bay
of Cadiz (SW Spain). Estuarine, Coastal and Shelf Science,
39: 127-141.
45
(2016)
Fe and Cu organic ligands in natural incubation experiments
A.G. Gonzalez1,2, G. Sarthou1,2, F. Chever1,2, F. Quéroué1,2, A. Bowie3, P. van der Merwe3, M.
Cheize1,2, M. Sirois5, E. Bucciarelli1,2
1
Université de Bretagne Occidentale, IUEM, 29200 Brest, France
LEMAR-UMR 6539, CNRS-UBO-IRD-IFREMER, Place Nicolas Copernic, 29280 Plouzané, France
3Antarctic Climate and Ecosystems CRC, University of Tasmania, Hobart, Tasmania, Australia
2
ABSTRACT
Trace metals, like Fe and Cu, can chemically compete in natural waters both inorganically and organically. This
interaction affects the speciation of metals in seawater. Both elements can also biologically compete, because of
the replacement in the photosynthetic apparatus of Fe-rich cytochrome C6 by Cu-containing plastocyanin and of
the use of the multi-copper oxidase in Fe transport systems of some phytoplankton species. Fe and Cu incubations
were carried out in natural seawater during the oceanographic cruise KEOPS 2 in the Southern Ocean, as a part of
GEOTRACES program, from 10th October to 20th November 2011 aboard the R.V. Marion Dufresne.
Experiments were performed with control experiments (no metal additions), Fe additions (1 nM Fe3+), Cu
additions (0.5 nM Cu2+) and Fe+Cu additions (1 nM Fe3+ +0.5 nM Cu2+), at three contrasted stations: The R-2
station, which in the High Nutrient-Low Chrorophyll (HNLC) area, the A3-1 station above the Kerguelen Plateau,
and the E-3 station in a permanent meander of the Polar Front, East of the Kerguelen Islands. The following
parameters were measured: concentrations of chlorophyll-a (Chl-a), biogenic silica (bSiO2), particulate organic
carbon and nitrogen (POC, PON), nitrate (NO3-), orthosilicic acid (Si(OH)4), flow cytometry, microscopy, Fe and
Cu organic speciation.
A clear gradient in Fe limitation was observed among the three stations, with R-2 showing a strong Fe-limition, E3 a mild-limitation, and A3-1 no limitation. The Cu addition did not show any significant effect in our
experiments.
INTRODUCTION
Fe is an essential micronutrient that limits primary
productivity in up to 40% of the global oceans [1]. Recent
studies demonstrated the evidence of Cu limitation for the
phytoplankton growth, especially under Fe limiting
conditions [2].
The interaction of Fe and Cu has been recently studied in
terms of redox interaction [3]. This interaction should
imply effects on the phytoplankton dynamic and the
possible production of organic ligands to control Fe and Cu
speciation in seawater. Note that Fe and Cu speciation is
dominated by the organic ligands (over 99%), increasing
their solubility and making them more bioavailable.
Microorganisms can modify the concentration and the type
of organic ligands in solution depending to the natural
conditions [4].
During the oceanographic cruise KEOPS 2 (PI: S. Blain,
Fig. 1) in the Southern Ocean, as part of the GEOTRACES
program, the Fe and Cu effects on the phytoplankton
dynamic was studied, measuring a number of parameters as
of chlorophyll-a (Chl-a), biogenic silica (bSiO2),
particulate organic carbon and nitrogen (POC, PON),
nitrate (NO3-), orthosilicic acid (Si(OH)4), flow cytometry,
microscopy, Fe and Cu organic speciation. These results
will help us to improve our knowledge about the natural
phytoplankton dynamic under different levels Fe, Cu and
Fe+Cu.
Fig. 1. Map of stations during KEOPS II.
MATERIAL AND METHODS
The study area was located at the vicinity of the Kerguelen
Islands in the Indian sector of the Southern Ocean where
phytoplankton blooms are observed each summer.
Experiments were carried out during the KEOPS 2 cruise
46
(2016)
from 10 October to 20 November 2011 aboard the R.V.
Marion Dufresne II (TAAF/IPEV). Experiments were
performed at three stations: R-2 (50.39°S/66.69°E)
considered as the reference station of the KEOPS 2
experiment [5], A3-1 (50.63°S/72.08°E) located above the
south Kerguelen plateau, and E-3 (48.70°S/71.97°E),
located in a recirculation area North East of the Kerguelen
Islands (Fig. 1).
Samples were collected using a trace metal clean rosette
(TMR, model 1018, General Oceanic). Seawater for the
incubations was collected in the surface mixed layer. All
the samples were manipulated in a clean container. The
bottles were acid-cleaned before use following the
GEOTRACES recommendations.
For each experiment and time point, duplicate bottles were
spiked with 1 nmol L-1 FeCl3 (+ Fe), or 0.5 nmol L-1
Cu(NO3)2 (+ Cu), or a combination of both (+ Fe + Cu). A
continuous running seawater system supplied water from
the sea surface, allowing the maintenance of in situ surface
temperature as well as the perpetual motion of the free
floating bottles that prevented any settling of material.
RESULTS AND DISCUSSION
Station R-2 is the reference stations. It is a typical High
Nutrient Low Chlorophyll (HNLC) station with low
dissolved Fe and Cu, Chl-a, BSi, POC and PON. The
concentration of dissolved Fe was higher at E-3 (0.38 nmol
L-1) and dissolved Cu was higher at A3-1 (1.93 nmol L-1).
The addition of Fe invokes a rapid response of the
phytoplankton community, at R-2 and E-3 stations. Chl-a
concentrations increase after additions of Fe and Fe+Cu,
achieving ~11 µmol L-1 after 14 days and 12 days,
respectively. The rest of parameters measured (BSi, POC,
PON and nutrients) followed the same trend in both
stations, with no significant effect over time for control and
+Cu additions, but with clear effect after +Fe and +Fe-Cu
treatment. At A3-1, the addition of metals did not represent
any significant increase in Chl-a, BSi, POC, PON and
nutrients.
The flow cytometry results showed that only
nanoeukaryotes were sensitive to the presence of Fe and Cu
in solution, showing significant increase of cell
concentration over incubation time. The most important
effect was also measured for stations R-2 and E-3, where
nanoeukaryotes increased after +Fe and +Fe-Cu between 2
and 3.5 times, respectively.
This dynamic change in stations should reveal effects on
the Fe and Cu organic speciation during the incubations.
Cu-organic ligands always slightly increased for the control
from the initial time and the last day of incubation. The
presence of Fe, Cu and Fe-Cu in solution never presented
clear trends. The higher Cu-ligand concentration was
measured for E-3 after +Fe treatment with 85 ± 2 nmol L-1.
However, this concentration was practically constant for all
the conditions at station A3-1 (43.8 ± 0.1 nmol L-1). The
Cu-ligand concentration increased from 28.2 ± 0.5 nmol L-1
(control) to 68.1 ± 0.3 nmol L-1 (+Fe+Cu) after 14 days of
incubations. These ligands can be ranked as weak Culigands according to the log Kcond value that was between
11.0 and 12.8 for most of the incubations. Only A3-1
showed strong ligands in the control and after +Cu and +Fe
treatment.
The concentration of Fe-ligands (0-3.76 nmol L-1) was
always lower compared to that for Cu-ligands (0-85 nmol
L-1). However, Fe-ligands increased after +Fe and +Fe+Cu
treatment for the stations R-2 (from 1.9 ± 0.3 to 2.5 ± 0.1
nmol L-1) and E-3 (from 1.7 ± 0.3 to 3.8 ± 0.2 nmol L-1). In
addition, Fe-ligands always decreased after +Cu additions
compared to the control for R-2 and E-3. The Fe-ligands
can be ranked as strong ligands, with log Kcond values
between 11.1 ± 0.3 and 12.6 ± 0.1.
ACKNOLEDGEMENTS
This research was supported by the Institut National des
Sciences de l’Univers (INSU), the French Polar Institute
(Institut Polaire Emile Victor, IPEV), the GIS EUROPOLE
MER, and the Agence National de la Recherche (ANR2010-BLAN-614). The PhD fellowship of FQ was cofunded by the University of Tasmania and the University of
Brest. We would like to thank the captain and the crew of
the R.V. Marion Dufresne, Stephane Blain and Bernard
Quéguiner as chief scientist and project coordinator of the
KEOPS 2 cruise.
REFERENCES
1 – Moore JK, Doney SC & Lindsay K, 2004. Upper ocean
ecosystem dynamics and iron cycling in a global threedimensional model. Global bioegeochim. Cycles,
18:GB4028.
2 – Peers G, Quesnel SA & Price NM, 2005. Copper
requirements for iron acquisition and growth of coastal and
oceanic diatoms. Limnol. Oceanogr., 50:1149-1148.
3 – Gonzalez AG, Perez-Almeida N, Santana-Casiano JM,
Millero FJ, Gonzalez-Davila M, 2016. Redox interactions
of Fe and Cu in seawater. Mar. Chem., 179:12-22.
4 – Rico M, Lopez A, Santana-Casiano JM, Gonzalez AG,
Gonzalez-Davila M, 2013. Variability of the phenolic
profile in the diatom Phaeodactylum tricornutum growing
under copper and iron stress. Limnol. Oceanogr., 58:144152.
5 – Blain S, Capparos J, Guéneuguès A, Obernosterer I &
Oriol L, 2015. Distributions and stoichiometry of dissolved
nitrogen and phosphorus in the iron-fertilized region near
Kerguelen (Southern Ocean). Biogeosciences, 12:623-635.
47
(2016)
Utilización de pigmentos para la caracterización del microfitoplancton el en
Golfo de Cádiz (2014-2015)
C. González-García 1, 2, *, C. García-Muñoz 2, M.C. González-Cabrera 3, M.P., Jiménez 3, Jesús
Forja 1 y L.M. Lubián 2
(1) Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz, Campus Universitario Río San Pedro, 11510 –
Puerto Real, Cádiz, Andalucía, España.
(2) Instituto de Ciencias Marinas de Andalucia (CSIC), Campus Universitario Río San Pedro, 11510 – Puerto Real, Cádiz,
Andalucia, España.
(3) Instituto Español de Oceanografía. Centro Oceanográfico de Cádiz. Puerto Pesquero, Muelle de Levante s/n. Apdo.
2609. E-11006 Cádiz, España
*Correo del autor: [email protected]
RESUMEN
En este trabajo se estudia la evolución de clorofila así como los pigmentos asociados al microfitoplancton en el
Golfo de Cádiz durante los años 2014 y 2015. El estudio se basa en los datos obtenidos a lo largo de tres radiales
del Golfo de Cádiz, la desembocadura del río Guadalquivir, el caño Sancti Petri y el cabo de Trafalgar. Se
observan importantes variaciones estacionales de clorofila, con concentraciones que varían desde máximos de
2.36 µg/L en primavera a mínimos de 0.02 µg/L en los muestreos más invernales. Se han determinados las
concentraciones de pigmentos asociados a la clorofila a, observando, también en ellos, una estacionalidad en
cuanto cantidad y diversidad. Se han determinado los principales pigmentos del microfitoplancton y se han hecho
relaciones con el total de clorofila a mediante el software CHEMTAX, con el cual se ha establecido al grupo de
las diatomeas como el más abundante en el Golfo de Cádiz a excepción de fenómenos de bloom y proliferaciones
locales.
INTRODUCCIÓN
El Golfo de Cádiz se encuentra al suroeste de la Península
ibérica, entre el cabo San Vicente, en Portugal, y el
Estrecho de Gibraltar. Está caracterizado por una
plataforma continental ancha, entre 50 y 15 kilómetros,
donde tienen lugar los procesos que mayor consecuencia
tienen en la abundancia y distribución de las poblaciones
fitoplanctónicas, como son los procesos de marea, el aporte
de nutrientes desde el continente o el efecto del viento
entre otros. Las masas de agua en el Golfo de Cádiz [1]
determinan variables físicas como temperatura y salinidad,
así como fenómenos de afloramiento y hundimiento.
Los mayores niveles de clorofila se dan en las zonas
costeras debido al aporte de nutrientes. La zona con mayor
concentración de clorofila durante los dos años analizados
es la desembocadura del río Guadalquivir, con valores
medios de 1.40 µg/L en periodos de mayor abundancia.
Los procesos de afloramiento en las cercanías del cabo
Trafalgar, producidos por efectos de la entrada de agua
mediterránea a través del estrecho [2], producen un
levantamiento de la nutriclina en este punto y favore la
subida hacía capas fóticas de células microfitoplantónicas
así como nutrientes hundidos en el mar Alborán [3]. Este
aumento en la disponibilidad de nutrientes tiene como
consecuencia un aumento ocasional de microfitoplanctón
en esta zona.
La determinación de pigmentos asociados a los diferentes
grupos del microfitoplancton, permite determinar los
grupos mayoritarios en estos procesos de afloramiento, así
como conocer los mayores grupos beneficiados de los
aportes de nutrientes. La cromatografía liquida de alta
eficacia (HPLC) no ha sido solo aplicada para la
determinación de la clorofila a, también para clorofilas,
xantofilas y carotenoides asociados [4]. El mejor método
para analizar los resultados de pigmentos y conocer su
contribución al total de clorofila a es el programa
informático CHEMTAX [5], pudiendo tener una visión
general de como los diferentes grupos de fitoplancton
participan en la distribución de clorofila a en la zona.
El objetivo de este trabajo es la caracterización de los
cambios de distribución, abundancia y biodiversidad de las
comunidades fitoplanctónicas en los años 2014 y 2015 en
el Golfo de Cádiz.
MATERIAL Y MÉTODOS
El área de estudio se localiza en el Golfo de Cádiz, al
suroeste de la península ibérica. Las muestras han sido
tomadas a lo largo de 8 campañas oceanográficas durante
48
(2016)
los años 2014 y 2015 dentro del proyecto STOCA (Series
Temporales de datos Oceanográficos en el Golfo de Cádiz)
que desarrolla el Instituto Español de Oceanografía (IEO).
En cada campaña se han realizados 3 transectos con un
total de 16 estaciones perpendiculares a la desembocadura
del río Guadalquivir y caño Sancti Petri, y al cabo de
Trafalgar.
Han sido tomadas muestras para clorofila fraccionada
mediante un sistema de doble filtración, determinando por
fluorimetría
la
fracción
correspondiente
al
microfitoplancton (>20µm) respecto al total de clorofila. Se
han tomado muestras superficiales y en el máximo de
fluorescencia profundo (DFM) para el análisis de
pigmentos mediante HPLC (Walters Alliance HPLC
System, con detector de fluorescencia). Para el procesado
de los datos obtenidos para pigmentos se ha utilizado el
software informático CHEMTAX que ayuda a comprender
que pigmentos pertenecen a cada grupo del
microfitoplancton y crear una proporción entre ellos
respecto la clorofila a total. Para la interpretación de los
datos se han utilizado los resultados de distribución y
concentración de nutrientes recogidos durante las
campañas en el Golfo de Cádiz.
RESULTADOS Y DISCUSIÓN
La distribución de la clorofila del rango de tamaño
microfitoplanctónico muestra sus máximos en las zonas
más costeras de las secciones situadas en el Guadalquivir y
en Sancti Petri. Generalmente coinciden con un mínimo en
la concentración de nutrientes, debido a su rápido consumo
y a la limitación de su entrada desde el fondo por una
termoclina permanente situada por debajo de la zona fótica.
También se han detectado máximos relativos de clorofila
en estaciones de la radial de Trafalgar. Aparecen de forma
ocasional, no parecen seguir un patrón estacional y están
relacionados con un descenso en la concentración de
nutrientes en las capas superficiales. Anteriores trabajos en
la zona cercana al estrecho [6, 7] describen este
afloramiento marcado por un aumento de la clorofila
debido a la mezcla vertical que experimenta las aguas
atlántica y mediterránea en el paso por el Umbral de
Camarinal. Esta mezcla supone un aporte de nutrientes
desde las capas más profundas a las más superficiales.
Basándose en trabajos anteriores [4], así como a los
pigmentos mayoritarios en las muestras se eligieron cuatro
grupos del microfitoplancton para la caracterización del
Golfo de Cádiz: haptofitas, dinofitas, prasinofitas y
bacilariofitas.
Estos
grupos
tienen
pigmentos
característicos que no comparten con el resto, por lo que es
fácil su identificación y posterior análisis con la
herramienta CHEMTAX. Durante las STOCA 2014 y
2015, las bacilariofitas fueron en grupo más importante en
todas las estaciones, seguidas de las haptofitas y dinofitas.
Las relaciones entre los cuatro grupos varían, viéndose
cambios según masas de agua, épocas del año, diferencias
costa-océano y afloramientos, como el de Trafalgar, en el
que se ve un cambio de proporción entre unos pigmentos
característicos de un grupo a otro, demostrando así que un
aporte de nutrientes adicional determina el crecimiento de
unos grupos respecto otros.
GD 0614
100
50
0
GD1 GD2 GD3 GD3 GD4 GD4 GD4 GD5 GD5 GD5 GD6 GD6 GD6 GD6
5m 5m 5m 33m 5m 31m 100m 5m 40m 50m 5m 85m 95m 400m
Prasinofitas
Bacilariofitas
Dinofitas
Haptofitas
Figura 1. Resultados del programa CHEMTAX
(abundancias pigmentarias respecto total de clorofila a)
correspondientes a la campaña STOCA junio 2014, radial
de Gualdaquivir. Predominio del grupo de las diatomeas en
todas las estaciones, seguido por el de las dinofitas.
AGRADECIMIENTOS
Agradecer a la tripulación de los B/O Ramón Margalef y
Ángeles Alvariño por su participación en las campañas.
Este trabajo ha sido financiado mediante el proyecto
STOCA del Instituto Español de Oceanografía y el
proyecto CTM2014-59244-C3.
REFERENCIAS
1 - Criado-Aldeanueva, F. et al., 2006. Distribution and
circulation of water masses in the Gulf of Cadiz from in
situ observations. Deep Sea Research Part II: Topical
Studies in Oceanography, 53(11-13), pp.1144–1160.
2 - Macías, D. et al., 2008. Chlorophyll maxima and water
mass interfaces: Tidally induced dynamics in the Strait of
Gibraltar. Deep-Sea Research Part I: Oceanographic
Research Papers, 55(7), pp.832–846.
3 - Prieto, L. et al., 1999. Phytoplankton, bacterioplankton
and nitrate reductase activity distribution in relation to
physical structure in the northern Alboran Sea and Gulf of
Cadiz (southern Iberian Peninsula). Boletin Instituto
Español de Oceanografia, 15, pp.401–411.
4 - Goela, P.C. et al., 2014. Using CHEMTAX to evaluate
seasonal and interannual dynamics of the phytoplankton
community off the South-west coast of Portugal. Estuarine,
Coastal and Shelf Science, 151, pp.112–123.
5 - Mackey, M.D., 1996. CHEMTAX- a program for
estimating class abundances from chemical markers :
application to HPLC measurements of phytoplankton.
Marine Ecology Progress Series, 144, pp.265–283.
6 - Ruiz, J. et al., 2001. Surface distribution of chlorophyll,
particles and gelbstoff in the Atlantic jet of the Alborán
Sea: From submesoscale to subinertial scales of variability.
Journal of Marine Systems, 29(1-4), pp.277–292.
7 - Echevarria, F. et al., 2002. Physical – biological
coupling in the Strait of Gibraltar. Deep Sea Research Part
II: Topical Studies in Oceanography, 49, pp.4115–4130.
49
(2016)
Spatial distribution and estuarine sources of dissolved organic matter export
to the coastal zone in the Gulf of Cádiz, Spain
Enrique González-Ortegón1, Francisco Baldó1, María J. Bellanco1, Ricardo F. Sánchez-Leal1,
María P. Jiménez1, J. Pedro Cañavate2, César Vilas2
1
Insituto Español de Oceanografía, Centro Oceanográfico de Cádiz, Puerto Pesquero, Muelle de Levante s/n, 11006 Cádiz.
Spain.
2
IFAPA Centro El Toruño, Camino de Tiro Pichón, 11500 El Puerto de Santa María, Cádiz, Spain.
ABSTRACT
Dissolved organic matter (DOM) is a major component of the organic matter transported to the coastal zone by
rivers. It controls ecosystem-level processes (e.g. food web) and constitutes an important pathway for nutrients
transport from land to coastal waters. We know that estuarine discharges affect the primary production and
nutrient composition in the adjacent coastal area. For instance, the current hypernutrification of the Guadalquivir
estuary may benefit primary production on adjacent coasts. However, studies on DOM in the Gulf of Cádiz waters
are unknown despite its importance in the global ocean functioning. The Gulf of Cádiz is under the estuarine
influence of three main estuaries: Guadiana, Tinto-Odiel and Guadalquivir. The present study evaluates the
relevance of DOM and the estuarine influence and environmental factors which determine its distribution in the
Gulf of Cádiz. Our results suggest that the Gulf of Cádiz water mass is receiving large amounts of dissolved
organic transported by the Guadiana and Guadalquivir rivers and much lesser from Tinto-Odiel. Thus, the
estuarine influenced area explained the fDOM variability in the Gulf of Cadiz and this variability was shaped by
turbidity, water depth and distance from the coast. Within the estuarine ecosystems, salinity and turbidity were the
main factors explaining the fDOM variability.
INTRODUCTION
Dissolved organic matter (DOM) influences aquatic
food webs and controls the availability of dissolved
nutrients and metals [1,2]. DOM is also important
from regional
and
global
biogeochemical
perspectives, as DOM constitutes an important
pathway for carbon (C), nitrogen (N), and
phosphorus (P) transport from land to sea (Harrison
2005). The quantification of these elements from
their sources in the continental shelf has been poorly
studied. The Gulf of Cádiz water masses are directly
influenced by the three main estuarine systems
(Guadiana, Tinto-Odiel and Guadalquivir). Nutrient
export to neighbouring coastal waters generates
phytoplankton blooms on the shelf [3]. In the
Guadalquivir estuary 17% and 83% of the estuarine
SPM concentration was organic and inorganic matter,
respectively [4]. However, at present we do not know
how fDOM varies through the Gulf of Cádiz and how
it is interacting with environmental variables. We
hypothesise that fDOM spatial variability in the GC
is determined by the rivers influence by transport to
costal waters. Advanced sensor technology is widely
used in aquatic monitoring and research. We used the
YSI EXO2 multiparametric sonde provided by six
sensors and an integral pressure transducer. Each
sensor measures its parameter via a variety of
electrochemical, optical, or physical detection
methods. We used the capability of water chemistry
sensors embedded in this new sensor platform to
document spatial variability in the Gulf of Cádiz.
This new sensor platform continuously samples the
mouth of the three main estuaries and water column
in the Gulf of Cádiz
MATERIALS AND METHODS
Sampling was carried out on board the Ramon Margalef
oceanographic vessel during March 2016 (Figure 1). We
used a YSI multiparameter sonde (EXO2; Yellow Springs,
OH). The EXO2 sonde uses a combination of electrical and
optical sensors for specific conductivity, water
temperature, pH, dissolved oxygen, turbidity, fluorescent
dissolved organic matter (fDOM), chlorophyll-a
fluorescence, and phycocyanin fluorescence. Physical
parameters of the EXO2 sonde were highly correlated
(R=0.8, p<0.01) with the CTD-ADCP. A multivariate
approach to spatial analysis among ecosystems (estuaries
vs. GC), within ecosystems estuaries and GC (Radials)
differences, and depth and coastal influenced area (distance
from the main land) differences in the total algae and
fDOM was followed using the PRIMER 6.1 (Plymouth
50
(2016)
Routines in Multivariate Ecological Research) computer
software pack. Multivariate data analysis was carried out
by non-metric multidimensional scaling (MDS) ordination
with the Euclidian distance similarity. Our physical dataset
includes temperature, salinity, dissolved oxygen, turbidity
and pH. We used generalised additive models (GAMs) to
test the physical factors effects on fDOM.
temperature and turbidity in all radials, while depth had a
negative effect, reducing fDOM concentration (Fig. 2). In
the estuaries, turbidity and salinity were found to have a
negative effect (Fig. 3).These results suggest that the
estuarine influence from the Guadalquivir and Guadiana
estuaries explains the dissolved organic matter variability
found in the Gulf of Cádiz. In general, fDOM
concentrations are much higher in estuaries than in the
open ocean, though concentrations are highly variable (2).
Although variations in fDOM are primarily the result of
natural processes, human activities such as freshwater
discharges and wetland drainage can affect the levels in
estuarine systems with carrying out effects on the Gulf of
Cádiz. Human transformations of the Tinto-Odiel estuary
would explain the high dissimilarity respect to the other
two estuaries. These novel observations resulted in highdensity, mesoscale spatial data and revealed unknown
variability in physical, chemical, and biological factors.
Figure 1. Sampling locations for the different oceanographic
cruises. Radials from right to left: TF Trafalgar, SP Sancti Petri,
GD Guadalquivir, TO Tinto y Odiel, and GU Guadiana.
l
Sa
We found an expected high dissimilarity among
ecosystems in terms of total algae and fDOM concentration
(ANOSIM analyses R=0.9, p<0.001). However, among
estuarine samples the Tinto-Odiel ones were closer to the
Gulf of Cádiz samples (low fDOM values) than the
Guadiana and Guadalquivir ones (the highest fDOM
values). Among the oceanic samples, the average similarity
was low (R=0.49, p<0.001) being this variability explained
mainly by the distance from the coastline and the radials
(Fig. 2). Most of the differences were found between the
station closer to the coastline (Stations 1) and the furthest
one (R = 0.6, p< 0.001).
tor
linear predic
RESULTS AND DISCUSSION
rio
tu a
fEs
Figure 3. fDOM prediction as a function of salinity in the
Guadalquivir (right) and Guadiana (left) estuaries. Tinto-Odiel has
not been showed due to the absence of a salinity gradient.
ACKNOWLODGEMENTS
This work was financed by the MICCIN grants DILEMA
(CTM2014-59244-C3-2-R).
REFERENCES
De
pth
ictor
linear pred
fRa
dia
l
Figure 2. fDOM prediction as a function of depth and the radials
of the Gulf of Cádiz salinity. Radials from right to left: TF, SP,
GD, TO and GU (see legend in the figure 1).
Nonparametric models (GAM) were fit to the data to
estimate the partial effects of the various covariates on
fDOM. In the Gulf of Cádiz, we found positive effects of
1 - Findlay, S. E. G., and R. L. Sinsabaugh (2003), Aquatic
Ecosystems: Interactivity of Dissolved Organic Matter, 512
pp., Elsevier, New York.
2- Harrison, J.A., N. Caraco, and S. P. Seitzinger (2005),
Global patterns and sources of dissolved organic matter
export to the coastal zone: Results from a spatially explicit,
global model, Global Biogeochem. Cycles, 19, GB4S04.
3- Prieto L, Navarro G, Rodríguez-Gálvez S, Huertas IE,
Naranjo JM,Ruiz J (2009) Oceanographic and
meteorological forcing of the pelagic ecosystem on the
Gulf of Cadiz shelf (SW Iberian Peninsula). Cont Shelf Res
29:2122–2137
4 -González-Ortegón E, Drake P. Effects of freshwater
inputs on the lower trophic levels of a temperate
estuary:physical, physiological or trophic forcing? Aquat
Sci 2012;74: 455–69
51
(2016)
Quantification of total carbohydrates in microalgae extracts
Isabel Guerra1, Idaira Jerez1, Argimiro Rivero1, Milagros Rico1, Miguel Suárez de Tangil1 &
Juan Luis Gómez-Pinchetti2
1
Grupo QUIMA- Instituto de Oceanografía y Cambio Global, Universidad de Las Palmas de Gran Canaria, Campus de Tafira,
35017 Las Palmas de Gran Canaria, Canary Islands, Spain.
2
Banco Español de Algas, Instituto de Oceanografía y Cambio Global, Universidad de Las Palmas de Gran Canaria, Muelle de
Taliarte s/n, 35214 Telde, Canary Islands, Spain
RESUMEN
Las microalgas son organismos fotosintéticos con requisitos relativamente simples para el crecimiento y
localizadas en hábitats diversos tales como aguas marinas, dulces, salobres, residuales o en el suelo, bajo un
amplio rango de temperaturas, pH y disponibilidad de nutrientes. Algunas microalgas presentan un alto contenido
en carbohidratos pudiendo tener éstos función biológica estructural (mayoritariamente en forma de celulosa y
polisacáridos solubles) o función de almacenamiento (principalmente en forma de almidón). En el desarrollo de la
investigación abordada en el presente artículo se analizan 11 cepas de microalgas aportadas por el Banco Español
de Algas (BEA).
ABSTRACT
Microalgae are photosynthetic organisms with relatively simple growth requirements and located in habitats such
as seawater, freshwater, brackish waste or on the floor, under a wide range of temperature, pH and nutrients
availability. Some microalgae have high carbohydrate content which can have structural biological function
(mostly in the form of cellulose and soluble polysaccharides) or storage function (e.g. starch). In the development
of the present research, 11 microalgae strains provided by the Spanish Bank Algae (BEA) were analized.
INTRODUCCIÓN
Las microalgas y las cianobacterias son un conjunto
heterogéneo
de
microorganismos
fotosintéticos
unicelulares, eucariotas las primeras y procariotas las
segundas. Se localizan en hábitats diversos tales como
aguas marinas, dulces, salobres, residuales o en el suelo,
bajo un amplio rango de temperaturas, pH y disponibilidad
de nutrientes; se les considera responsables de la
producción del 50% del oxígeno y de la fijación del 50%
del carbono en el planeta.
Algunas microalgas pueden contener una gran cantidad de
carbohidratos acumulados en plastos como materiales de
reserva (mayoritariamente en forma de almidón) o como
componente principal de la pared celular (en forma de
celulosa y polisacáridos solubles) [1]. La acumulación de
carbohidratos en microalgas se debe a la fijación del
dióxido de carbono durante el proceso fotosintético. La
fotosíntesis es un proceso biológico que utiliza
ATP/NADPH para fijar y convertir el CO2, capturado
desde el aire, produciendo glucosa y otros azúcares a través
de una ruta metabólica conocida como el ciclo de Calvin.
Los carbohidratos contenidos en microalgas son complejos
y consisten en una mezcla de azúcares neutros, amino
azúcares y ácidos urónicos cuyas composiciones varían
entre especies y condiciones de crecimiento [2]. La
caracterización precisa de estos carbohidratos es
actualmente uno de los principales obstáculos para el
análisis de la composición detallada de las microalgas y de
los compuestos que exudan al exterior.
Los análisis más comunes de carbohidratos implican un
procedimiento de hidrólisis (ácida o alcalina) para romper
los polímeros en sus constituyentes monoméricos. En esta
investigación, los carbohidratos totales presentes en las
muestras de biomasa de alga se cuantifican mediante un
método analítico que utiliza antrona en presencia de ácido
sulfúrico que reacciona para formar un derivado del furano
de color verde (furfural o hidroximetilfurfural).
Seguidamente mediante espectrofotometría se determina la
concentración de carbohidratos presentes en la muestra ya
que ésta es función de la intensidad de color.
MATERIAL Y MÉTODOS
Se ha dispuesto de 11 cepas diferentes de microalgas y
cianobacterias procedentes del Banco Español de Algas
(BEA): (1) BEA0536B, Ankistrodesmus sp.; (2)
BEA0762B, Phormidiochaete sp.; (3) BEA0854B,
Nodularia spumigena; (4) BEA0031B, Chloroidium
saccharophilum; (5) BEA0190B, Pseudopediastrum
boryanum; (6) BEA0204B, Cosmarium blyttii; (7)
BEA0208B,
Cosmarium
sp.;
(8)
BEA0659B,
Pseudopediastrum boryanum; (9) BEA0666B, Spyrogyra
52
(2016)
sp.; (10) BEA0860B, Ochrosphaera verrucosa; (11)
BEA0990B, Chloromonas cf, reticulata.
Preparación de los extractos de microalgas
Se prepararon extractos acuosos mezclando 10 mg de cada
microalga previamente liofilizada con 10 mL de agua
destilada. Se someten a agitación durante 10 minutos en un
agitador magnético y posteriormente se introducen en un
equipo de ultrasonidos unos segundos. Tras 5 minutos
centrifugando se extrae el sobrenadante y es filtrado con un
filtro de jeringa de 0,45 μm.
Determinación de carbohidratos
extractos. El máximo contenido de carbohidratos se
encontró en las cepas Ochrosphaera verrucosa y Nodularia
spumigena, siendo la cepa Phormidiochaete sp. la que
ofrece menor contenido de carbohidratos.
Como conclusión, estos resultados indican que las especies
de microalgas analizadas pueden ser utilizadas para la
producción de bioetanol por fermentación de carbohidratos
y para la generación de biogás a partir de biomasa de
microalgas, principalmente las especies Ochrosphaera
verrucosa y Nodularia spumigena, que mostraron el mayor
contenido de carbohidratos.
AGRADECIMIENTOS
El contenido de carbohidratos se determinó mediante un
método descrito anteriormente que utiliza el reactivo de
antrona. Para ello, se preparó el reactivo pesando 0,1 g de
antrona y diluyendo con ácido sulfúrico al 98% hasta un
volumen de 50 mL. Se introdujo 1 mL de cada extracto de
microalga en tubos de ensayo y, a continuación, se
añadieron 2 mL del reactivo de antrona previamente
preparado. Durante 20 segundos se agitó la mezcla en un
equipo de agitación vórtex y, seguidamente, se introdujeron
los tubos en un baño de agua fría durante 2 minutos y en un
baño a 100°C durante 10 minutos. Finalmente, se dejó
atemperar los tubos de ensayo otros 10 minutos en un baño
de agua fría y se determinó la absorbancia a una longitud
de onda de 625 nm, frente a un blanco preparado con 1 mL
de agua destilada y 2 mL de reactivo de antrona tratado de
la misma forma. La cantidad de carbohidratos en las
muestras analizadas se determinó utilizando como
referencia una recta de calibrado obtenida a partir de
diferentes patrones de (+D)-Glucosa tratados con el mismo
procedimiento descrito previamente para las muestras
algales.
Los autores desean expresar su agradecimiento al Banco
Español de Algas (Taliarte, España) por la colección de
cepas que han aportado y que han hecho posible este
trabajo.
REFERENCIAS
1 – Chun-Yen hen, 2013. Microalgae-based carbohydrates
for biofuel production. Biochem. Eng. J., 78:1-10.
2 – David W. Templeton, 2012. Separation and
quantification of microalgal carbohydrates. J. Chromatogr.
A., 1270:225-234.
RESULTADOS Y DISCUSIÓN
La curva de calibrado fue obtenida para el rango de
concentraciones comprendidos entre 5 y 125 mg/L, siendo
realizados y analizados todos los patrones por triplicado.
Absorbancia
2.000
1.500
y = 0.0139x + 0.0265
R² = 0.9998
1.000
0.500
0.000
0
50
100
150
Concentración (mg/l)
Fig. 1. Curva de calibrado (+D)-Glucosa.
Sustituyendo los valores de absorbancias obtenidas para los
extractos de algas en la ecuación de la recta de calibrado
obtenida (y = 0,0139x + 0,0265), se ha cuantificado la
cantidad de carbohidratos totales contenidos en dichos
53
(2016)
Quantification of phenolic compounds in microalgae
Idaira Jerez-Martel1, Sara García-Poza1, Gara Rodríguez-Martel1, Cristina Afonso-Olivares1,
Milagros Rico1, Miguel Suárez de Tangil1 & Juan Luis Gómez-Pinchetti2
1
Grupo QUIMA, Instituto de Oceanografía y Cambio Global, Universidad de Las Palmas de Gran Canaria, Campus de
Tafira, 35017 Las Palmas de Gran Canaria, Canary Islands, Spain.
2
Banco Español de Algas, Instituto de Oceanografía y Cambio Global, Universidad de Las Palmas de Gran Canaria, Muelle
de Taliarte s/n, 35214 Telde, Canary Islands, Spain.
ABSTRACT
The extracts of several microalgae from the culture collection at the Spanish Bank of Algae (Ankistrodesmus sp.,
Spirogyra sp., Euglena cf. cantabrica and Caespitella pascheri) were screened for their radical scavenging
activity against the stable radical 1,1-diphenyl-2-picrylhydrazyl (DPPH). In addition, their phenolic profiles were
determined by using reversed phase high performance liquid chromatography (RP-HPLC), which allowed the
identification of 6 phenolic constituents: gallic acid, (+) catechin, (-) epicatechin, syringic acid, protocatechuic
acid and chlorogenic acid. Microalgae Euglena cf. cantabrica, Spirogyra sp. and Ankistrodesmus sp. showed the
presence of the phenolic compounds, having Euglena cf. cantabrica the highest amount (5.15 mg of gallic acid
and 1.26 mg of protocatechuic acid per gram of freeze-dried microalgae). Syringic acid was not detected and (+)
catequin, chlorogenic acid and (-) epicatechin were also quantified in Euglena cf. cantabrica (71.4, 77.9 and
7.09 µg per gram of dry material, respectively, two orders of magnitude lower than gallic and protocatechuic
acids). Besides, microalgae Euglena cf. cantabrica exerted the highest antioxidant activity, which may be related
to the presence of these compounds in high concentrations.
INTRODUCTION
Primary production is the ultimate source of organic matter
in the sea, but living biomass forms less than 1% of total
organic carbon in seawater, while more than 90% of
organic carbon occurs as non-living dissolved organic
carbon. In contrast to the organic reservoirs on land, the
processes by which dissolved organic matter (DOM) has
been formed are unclear, and actual sources and the
chemical nature of DOM are not well known [1]. DOM has
various functions and plays important roles in chemical,
biological and even physical oceanography. Despite DOM
composition strongly influences its role in the environment,
DOM characterization is still not routinely included in
many biogeochemical studies. DOM interacts with trace
metals and controls their dynamics.
Wells et al. [2] demonstrated that diatoms of the genus
Pseudo-nitzschia have adapted to iron limitation through
the production of a strong iron-complexing organic ligand,
domoic acid. The influences of Cu(II) and Fe(III) metals on
the cells and exudate of phenolic profiles of the green
microalgae
Dunaliella
tertiolecta
and
diatom
Phaeodactylum tricornutum have been demonstrated in our
previously reported studies focused on the implications of
polyphenols in microalgae growing under metal stress [3,
4]. Cells exposed to copper excreted a larger amount of
polyphenols as a protective mechanism to alleviate the
toxicity of copper in the solution. These phenolic
compounds are implicated countering metal toxicity at the
membrane surface and slowing down the toxicity of metals
in the extra-cellular media. In addition, phenolic
compounds exuded from microalgae, such as sinapic acid
and (+) catechin, have an influence in iron redox chemistry
by favouring the persistence of Fe(II) for their
requirements [5]. Despite the important role of phenolic
compounds, few reports have focused on the identification
and quantification of polyphenols in microalgae [3,4,5].
The main objective of this work was to identify and
quantify the following 6 phenolic compounds in
microalgae extracts: gallic acid (GA), (+) catechin (C), (-)
epicatechin (E), syringic acid (SA), protocatechuic acid
(PA) and chlorogenic acid (CA). Microalgae strains,
bioprospected at the Canarian region and deposited at the
culture collection of the Spanish Bank of Algae
(Taliarte, Spain), were: Ankistrodesmus sp., Spirogyra sp.,
Euglena cf. cantabrica and Caespitella pascheri. The
antioxidant activities of extracts were also determined with
regard to their potential application for particular food and
pharmaceutical purposes. The increased demand for
healthy foods could find a nontraditional ally in
microalgae.
54
(2016)
MATERIAL AND METHODS
Microalgae strains were provided by the Spanish Bank of
Algae (Taliarte, Spain). Chromatographic analysis was
performed on a Liquid Chromatography Varian system
equipped with a diode array detector (DAD) and connected
to a Star software. The Radical Scavenging Activity (RSA)
was determined by measuring the loss of DPPH color at
515 nm after reaction with the samples.
RESULTS AND DISCUSSION
Yields of several microalgae extractions were evaluated
and the results are presented in Figure 1. Increasing
extraction efficiency was found when water was used as
extracting solvent with two exceptions: Euglena cf.
cantabrica and Caespitella pascheri, which showed higher
yield of extraction by using methanol. The highest yields
were found in microalgae Euglena cf. cantabrica (57.9%).
80
60
40
Methanol
57.9
25.7
20
0
Euglena cf.
cantabrica
Water
30.2
16.3 13.7
12.8 13.2
Caespitella
pascheri
Spirogyra sp.
18.3
Ankistrodesmus
sp.
Fig 1. Yield of extractions expressed as a percentage by
weight of the freeze-dried starting material.
The strongest DPPH radical scavenging capacity was
associated to Euglena cf. cantabrica extracts (100%
inhibition) with a half-life (t1/2 = time required for reducing
initial concentration of DPPH by 50%) lower than 2.1
seconds, followed by methanol extract of Spirogyra sp.
(61.6%) with a t1/2 of 202 seconds (when the extracts were
prepared by mixing 10 mg of microalgae biomass per mL).
At the same proportion of 1 mg mL-1, Euglena cf.
cantabrica gave higher activity (71%) in inhibiting DPPH
radical than BHT (26%) and lower activity and t1/2 (4
seconds) than BHA (91% with a t1/2 of 130 seconds), being
BHA and BHT synthetic preservatives commonly used in
the food industry.
The presence of polyphenols in the extracts was confirmed
by comparing retention times and overlapping UV spectra
with those of the standard compounds. Among all tested
algae, Caespitella pascheri did not provide polyphenols.
However, GA was identified in Spirogyra sp. and PA was
detected in Ankistrodesmus sp. Microalgae Euglena cf.
cantabrica showed much higher amounts of GA and PA in
comparison with the other microalgae in the present study
(5.15 and 1.26 mg per gram of freeze-dried material,
respectively) and also presented relevant quantities of C,
CA and E.
As a conclusion, a selected group of phenolic compounds
was detected in several microalgae strains and may be
involved in microalgae cellular response to ROS, being
potential components of DOM.
ACKNOWLEDGMENTS
The authors would like to express their gratitude to José
Juan Santana Rodríguez for allowing the use of the HPLC
equipment and to the Spanish Bank of Algae for
providing microalgae strains. This research was
supported by the sponsors “Familia Megías Martínez” and
“Satocán Group” (Innova Program 2020 - Fundación
Universitaria de Las Palmas).
REFERENCES
1 - Ogawa H & Tanoue E, 2003. Dissolved Organic Matter
in Oceanic Waters. J. Oceanogr., 59:129-147.
2 - Wells ML, Trick CG, Cochlan WP, Hughes MP &
Trainer VL, 2005. Domoic acid: The synergy of iron,
copper, and the toxicity of diatoms. Limnol. Oceanogr.,
50(6):1908-1917.
3 - López A, Rico M, Santana-Casiano JM, GonzálezDávila M & González AG, 2015. Phenolic profile of
Dunaliella tertiolecta growing under high levels of copper
and iron. Environ. Sci. Pollut. Res., 22:14820–14828.
4 - Rico M, López A, Santana-Casiano JM, González AG
& González-Dávila M, 2013. Variability of the phenolic
profile in the diatom Phaeodactylum tricornutum growing
under copper and iron stress. Limnol. Oceanogr.,
58(1):144-152.
5 - Santana-Casiano JM, González-Dávila M, González
AG, Rico M, López A & Martel A, 2014. Characterization
of polyphenol exudates from Phaeodactylum tricornutum
and their effects on the chemistry of Fe(II)-Fe(III). Mar.
Chem.,
158:10-16.
55
(2016)
Estacionalidad de los flujos de CO2 agua -atmósfera en el Golfo de Cádiz
(2014-2015)
D. Jiménez-López1, S. Garrido1, N. Hernández-Poyuelo1, A. Sierra1, T. Ortega1, R. Ponce1,
M.J. Bellanco2, R. Sánchez-Leal2, A. Gómez-Parra1 y J. Forja1, *
1
Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz, Campus Universitario Río San Pedro, 11510 – Puerto
Real, Cádiz (España).
2
Instituto Español de Oceanografía. Centro Oceanográfico de Cádiz. Puerto Pesquero, Muelle de Levante s/n. Apdo. 2609. E11006, Cádiz (España).
*Correo del autor: [email protected]
RESUMEN
Se presentan las variaciones espacio-temporales de la presión parcial de CO2 en la zona nororiental de la
plataforma continental del Golfo de Cádiz. La base de datos corresponde a un total de 6 campañas oceanográficas
realizadas en 2014 y 2015. Los valores medios de pCO2 presentan importantes variaciones estacionales, con
intervalos comprendidos entre 340 – 380 µatm en invierno y primavera, y 425 – 445 µatm en situaciones de
verano y otoño. Las distribuciones de pCO2 no son homogéneas, presentando generalmente una disminución con
la distancia a la costa, así como asociados a eventos de elevada producción primaria (elevadas concentraciones de
clorofila y utilizaciones aparentes de oxígeno negativas). Los flujos con la atmósfera presentar valores negativos
durante el invierno y primavera, y positivos en verano y otoño, con un valor medio anual de -0.6 mmol m-2 d-1.
Por tanto, el Golfo de Cádiz en su conjunto actúa como un sumidero de CO2, con una capacidad de captación de
9.6⋅108 mol año-1.
INTRODUCCIÓN
Los océanos en su conjunto regulan el clima del Planeta
mediante un intercambio continuo de calor y gases de
efecto invernadero con la atmósfera [1]. De esta forma, se
estima que los océanos en su conjunto son capaces de
captar aproximadamente el 40 % de las emisiones
antropogénicas de CO2 a la atmósfera ([2], [3]). Sin
embargo, existe un intenso debate científico del papel que
desempeña los márgenes continentales en el ciclo global
del carbono, en gran parte debido en la definición de sus
límites. Como zona de transición entre los sistemas
costeros y los océanos, el intercambio de CO2 con la
atmósfera se encuentra condicionado por la distancia a
costa. De esta forma, la zona costera actúa generalmente
como fuente de CO2 a la atmósfera, mientras que las
plataformas continentales constituyen un sumidero de CO2
atmosférico, con una capacidad de captación de 0.3 - 0.4
Pg C año-1, lo que supone aproximadamente un 30% la
captación de CO2 por el océano global [4]).
Existen pocos estudios sobre la variabilidad de los flujos de
CO2 agua-atmósfera en el Golfo de Cádiz. Ribas et al. [5]
destacan la dependencia de sus valores con la distancia a
costa, de forma que en las zonas más someras, debido a los
aportes fluviales y a la intensificación de los procesos de
mineralización de la materia orgánica en lo sedimentos,
actúan como una fuente de CO2. Por otra parte, los flujos
de CO2 agua-atmósfera en la zona estudiada presenta una
intensa variación estacional, de forma que durante los
meses de Noviembre a Marzo actúa como un sumidero de
CO2, y en los meses de verano como una fuente. El balance
neto del intercambio de CO2 para todo el año muestra que
la zona nororiental del Golfo de Cádiz actúa como un
sumidero de CO2 atmosférico, con un flujo medio de -0,2
mmol m-2 día-1. Adicionalmente, De la Paz et al. [6]
describen los elevados valores de pCO2 que se detectan en
la desembocadura del Guadalquivir, que actúa como una
fuente de CO2 a la atmósfera, con flujos medios de 85
mmol m-2 día-1.
MATERIAL Y MÉTODOS
Las campañas se realizaron a bordo de los B/O Ángeles
Alvariño y Ramón Margalef al comienzo de las estaciones
climatológicas comprendidas entre primavera de 2004 y
verano de 2005.
La presión parcial de CO2 se ha registrado on line mediante
un sistema conectado a la toma de agua del continuo del
barco y que está provisto de un equilibrador mixto (tipo
burbuja” y “flujo laminar y un IRGA (LI-Cor 6262). Se ha
integrado en la base de datos la información del
termosalinógrafo y la estación meteorológica del barco.
Adicionalmente se han tomado muestras discretas de las
aguas superficiales para la cuantificación de oxígeno, pH,
56
(2016)
alcalinidad total, clorofila y nutrientes. Estas variables, que
permiten una mejor interpretación de la influencia de los
procesos biológicos sobre los valores de pCO2, se han
determinado utilizando los métodos habituales en
oceanografía.
RESULTADOS Y DISCUSIÓN
En la figura 1 se muestra las variaciones típicas de pCO2
encontradas en las aguas superficiales. En general se
observa como en las estaciones más cálidas (verano y
otoño) presentan valores entre 70 y 90 µatm superiores a
los encontrados en invierno y primavera, así como un
aumento con la cercanía a costa.
Tabla 1. Valores medios de temperatura, pCO2, velocidad
de viento (V) y flujo de CO2 agua-atmósfera (valor
negativo indicada captación y positivo emisión).
Mar 2004
Jun 2004
Oct 2004
Dic 2004
Mar 2005
Jun 2005
t
(°C)
15.5
21.1
21.7
18.2
15.7
20.9
pCO2
(µatm)
393.4
410.7
408.6
381.6
364.4
400.2
V
(m s-1)
7.0
6.1
6.3
7.3
4.7
6.9
Flujo
(mmol m-2 s-1)
- 0.4
1.1
1.0
- 2.5
- 2.8
0.1
El flujo medio para el periodo de tiempo estudiado es -0.6
mmol m-2 d-1, y por tanto, el Golfo de Cádiz actúa como un
sumidero de CO2, con una capacidad de captación
aproximada de 9.6⋅108 mol año-1 si se considera una
superficie estudiada de 4380 km2.
AGRADECIMIENTOS
Este trabajo ha sido financiado por los proyectos STOCA
(Instituto Español de Oceanografía) y CTM2014-59244C3.
REFERENCIAS
Fig. 1. Variaciones de pCO2 (µatm) en el Golfo de Cádiz
correspondientes a las campañas de diciembre de 2014 y
junio de 2015
En la tabla 1, donde se presentan los valores medios de
pCO2, velocidad de viento y flujo con la atmósfera para
cada campaña, puede apreciarse esta evolución estacional.
Cabe destacar la elevada temperatura media registrada en
octubre de 2004, superior incluso a las encontradas en
verano, así como la anomalía térmica de la campaña
realizada en invierno de 2004. Existe una cierta correlación
lineal entre los valores medios de pCO2 y la temperatura (r2
= 0.64), así como evidencias locales de la influencia de
procesos biológicos sobre pCO2. De esta forma, mínimos
relativos de pCO2 se ha asociado a altas concentración de
clorofila, pH elevados o valores negativos de la utilización
aparente de oxígeno, fundamentalmente en la zona afectada
por la descarga de nutrientes del Guadalquivir.
Takahashi et al. [7] propone un algoritmo para evaluar la
importancia relativa de los procesos térmicos y biológicos
sobre los valores de pCO2. Para el conjunto de valores
medios se obtiene un T/B de 1.12, indicando que son
fundamentalmente los procesos térmicos los responsables d
de la variación de pCO2 en el Golfo de Cádiz.
1 - Sarmiento, J. L. y Gruber, N. 2006. Ocean
Biogeochemical Dynamics. Princeton University Press,
528 pp.
2 - Tans, P.P., Fung, I.Y. y Takahashi, T., 1990.
Observational constraints on the global atmospheric CO2
budget. Science, 247: 1431-1438.
3 - Broecker, W.S. y Peng, T.H., 1992. Interhemispheric
transport of carbon dioxide by ocean circulation. Nature,
356: 587-589.
4 - Chen, C.T.A., y Borges, A.V., 2009. Reconciling
opposing views on carbon cycling in the coastal ocean:
Continental shelves as sinks and near-shore ecosystems as
sources of atmospheric CO2. Deep sea Research Part II:
Topical Studies in Oceanography, 56(8-10): 578-590.
5 - Ribas, M., Gómez-Parra, A. y Forja, J.M. 2011. Air-sea
CO2 fluxes in the north-eastern shelf of the Gulf of Cádiz
(southwest Iberian Peninsula). Mar. Chem 123 (1–4): 5666.
6 - De la Paz, M., Gómez-Parra, A., Forja, J.M., 2007.
Inorganic carbon dynamic and air-water CO2 exchange in
the Guadalquivir Estuary (SW Iberian Peninsula). Journal
of Marine Systems, 66(1-2): 265-277.
7 - Takahashi, T., Sutherland, S.C., Sweeney, C., Poisson,
A.,Metzl, N., Tilbrook, B., Bates, N.,Wanninkhof, R.,
Feely, R.A., Sabine, C., Olafsson, J., Nojiri, Y., 2002.
Global sea–air CO2 flux based on climatological surface
ocean pCO2, and seasonal biological and temperature
effects. Deep‐Sea Research II 49, 1601–1622.
57
(2016)
Influencia de los aportes costeros en la dinámica del CH4 en el Golfo de
Cádiz
D. Jiménez-López1, *, A. Sierra1, T. Ortega1, R. Ponce1, M.J. Bellanco2, R. Sánchez-Leal2, A.
Gómez-Parra1 y J. Forja1
1
Dpto. Química-Física. CACYTMAR.Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz, Campus
Universitario Río San Pedro, 11510 – Puerto Real, Cádiz, Andalucía, España.
2
Instituto Español de Oceanografía. Centro Oceanográfico de Cádiz. Puerto Pesquero, Muelle de Levante s/n. Apdo. 2609. E11006, Cádiz (España).
*Correo del autor: [email protected]
RESUMEN
Se han realizado muestreos en el Golfo de Cádiz durante 2014 y 2015 en los que se ha determinado la
concentración de CH4 lo largo de varias secciones: Guadalquivir, Sancti Petri y Trafalgar. El CH4 se ha medido
utilizando un cromatógrafo de gases. Se ha observado un aumento de CH4 en zonas profundas como consecuencia
de las características hidrodinámicas de la zona, así como por los procesos de remineralización bentónica
ocurridos en el sedimento. Las concentraciones más elevadas se han detectado cerca de costa debido a los aportes
continentales. Los mayores flujos a la atmósfera se han estimado en los meses estivales debido a las altas
temperaturas alcanzadas durante esas campañas. Toda la zona estudiada se comporta como una fuente de CH4 a la
atmósfera, con emisiones globales de 0,61 y 0,75 Gg CH4 año-1 en 2014 y 2015 respectivamente.
INTRODUCCIÓN
El metano es responsable de aproximadamente el 20% del
efecto invernadero, con un efecto 25 veces mayor que el
CO2. Dicho gas tiene origen natural (ej. los humedales) y
antropogénico (ej. agricultura), siendo este segundo origen
responsable de más de la mitad de las emisiones actuales
de CH4 [1].
El CH4 se forma durante la descomposición de la materia
orgánica mediante el proceso anaeróbico de la
metanogénesis.
Las
principales
reacciones
de
metanogénesis son la fermentación de la materia orgánica y
la reducción del CO2 [2]. Además, el metano en el medio
marino puede tener un origen no biogénico ya sea por la
filtración de CH4 termogénico, mediante estructuras
geológicas como los volcanes de fango o mediante la
disolución de hidratos de gas [3].
MATERIAL Y MÉTODOS
El muestreo se realizó en la parte oriental del Golfo de
Cádiz, situado al suroeste de la península Ibérica. La
circulación en el Golfo de Cádiz se encuentra dominada
por el intercambio de masas de aguas del Atlántico y el
Mediterráneo a través del Estrecho de Gibraltar. Además
del flujo de agua Mediterránea, el golfo de Cádiz recibe
considerables aportes de ríos que desembocan en la cuenca,
como el Guadalquivir.
Las muestras se recogieron en tres transectos
perpendiculares a costa a diferentes profundidades (Fig. 1),
durante las campañas STOCA 2014 y 2015
correspondientes a las cuatro estaciones del año, a bordo de
los buques Ángeles Alvariño y Ramón Margalef.
Fig. 1. Transectos de la zona de estudio: Guadalquivir
(GD), Sancti Petri (SP) y Trafalgar (TF).
Para el análisis de CH4, las muestras se tomaron por
duplicado en frascos Winkler de 250 mL, se fijaron con
HgCl2 para inhibir procesos microbiológicos, y se sellaron
con grasa Apiezon® para prevenir el intercambio gaseoso
con la atmósfera.
Las medidas de CH4 disuelto se realizaron utilizando un
cromatógrafo de gases Bruker® GC-450 provisto de un
detector de ionización de llama, tomando unos 25 g (±0,01
g) de la muestra mediante el uso de una jeringa de cristal
(Agilent P/N 5190-1547) y 25 mL de un gas patrón de
concentración conocida (1800 ppbv). Esta operación se
realizó por duplicado para cada frasco Winkler. Tras esto,
se agita la jeringa durante 5 minutos (VIBROMATIC
Selecta) y se deja reposar para alcanzar una situación de
equilibrio. Por último, el gas es inyectado en el
cromatógrafo de gases.
La concentración de gases en el agua se calculó a través de
las medidas realizadas sobre el espacio de cabeza de las
58
(2016)
muestras, usando las solubilidades propuestas por
Wiesenburg & Guinasso (1979) [4].
Para la estimación de los flujos de gases en la interfase
atmósfera-océano en la zona de estudio se utilizó la
siguiente expresión:
F = k(CW – C*)
-1
donde k (cm h ) es la velocidad de transferencia del gas,
Cw (mol L-1) es la concentración del gas en el agua, y C*
(mol L-1) es la solubilidad del gas a la temperatura de
equilibración (25 ± 1 °C) y a la salinidad de la muestra. Un
flujo positivo indica transferencia del gas del agua a la
atmósfera
RESULTADOS Y DISCUSIÓN
Las concentraciones medidas en este estudio presentan
variabilidad estacional, con valores medios más elevados
durante otoño (12,53 ± 1,84 nM) y más bajos en primavera
(8,08 ± 0,66 nM). Estos valores son más elevados que los
encontrados por Ferrón et al (2010) [5] en aguas del Golfo
de Cádiz durante estas estaciones del año. En todas las
estaciones estudiadas, las aguas superficiales se encuentran
sobresaturadas de CH4. Los valores coindicen con
concentraciones de CH4 obtenidas en el estuario de
Changjiang y su área marina adyacente [6]. Sin embargo,
son superiores a los datos encontrados en zonas oceánicas
[7].
En los transectos de Guadalquivir y Sancti Petri se observa
claramente que los valores más elevados están asociados a
aguas costeras, con valores máximos que varían entre
10,00 y 19,65 nM en Guadalquivir en las diferentes
estaciones del año, y valores entre 10,82 y 20,22 nM en
Sancti Petri dependiendo de la campaña. En el transecto de
Trafalgar también se encuentran altos valores en el fondo
de las estaciones más costeras, con concentraciones que
varían entre 10,08 y 17,36 nM. Las plataformas
continentales y los estuarios son los responsables de
aproximadamente el 75% de las emisiones globales
oceánicas de CH4 [8], en concordancia con las mayores
concentraciones de CH4 medidas en este estudio en las
zonas más costeras. De hecho, la zona costera de las
secciones de Guadalquivir y Sancti Petri reciben aportes de
materia orgánica provenientes del río Guadalquivir en el
primer caso, y de la red de caños y marismas en el caso de
Sancti Petri, incorporándose una parte importante de esta
materia orgánica a los sedimentos. Por tanto, los altos
valores de CH4 encontrados en el fondo se deben a la
metanogénesis producida en el sedimento (Fig. 2).
Fig. 2. Transecto de Sancti Petri durante junio de 2015.
Al igual que las concentraciones de CH4, los flujos de este
gas presentan variabilidad estacional, encontrándose, en
general, los valores más elevados en otoño y verano, y los
más bajos a comienzos de primavera, donde se registraron
las temperaturas más bajas. Los flujos de CH4 máximos se
deben a las elevadas temperaturas medidas durante las
campañas estivales, lo que afecta a la transferencia del gas
a la atmósfera al disminuir su solubilidad.
Los flujos medios de CH4 son positivos, es decir, el Golfo
de Cádiz actúa como fuente de estos gases a la atmósfera.
Las emisiones globales del sistema son de 0,61 y 0,75 Gg
CH4 año-1 en 2014 y 2015 respectivamente para el área de
estudio (43,83 x 102 Km2). Este valor es superior al
calculado por Ferrón et al (2010) [5], 0,08 Gg año-1 aunque
para una superficie menor del Golfo de Cádiz (15,86 x 102
Km2).
AGRADECIMIENTOS
Este trabajo ha sido financiado por los proyectos STOCA
(Instituto Español de Oceanografía) y CTM2014-59244C3.
REFERENCIAS
1 - Intergovernmental Panel of Climate Change (IPCC),
2013. Climate Change 2013: The Physical Science Basis.
Contribution of Working Group I to the Fifth Assessment
Report of the IPCC. [Stocker, T.F., D. Qin, G.-K. Plattner,
M. Tignor, S.K. Allen, J. Boschung, A. Nauels, Y. Xia, V.
Bex & P.M. Midgley (eds.)]. Cambridge University Press,
Cambridge, United Kingdom and New York, NY, USA,
1535 pp.
2 - Reeburgh, W. S., 2007. Oceanic methane
biogeochemistry. Chemical Reviews, 107(2): 486-513.
3 - Judd, A. G., Hovland, M., Dimitrov, L. I., Garcia Gil,
S., & Jukes, V., 2002. The geological methane budget at
continental margins and its influence on climate change.
Geofluids, 2(2): 109-126.
4 - Wiesenburg, D. A., & Guinasso Jr, N. L., 1979.
Equilibrium solubilities of methane, carbon monoxide, and
hydrogen in water and sea water. Journal of Chemical and
Engineering Data, 24(4): 356-360.
5 - Ferrón, S., Ortega, T., & Forja, J. M., 2010. Temporal
and spatial variability of methane in the north-eastern shelf
of the Gulf of Cádiz (SW Iberian Peninsula). Journal of
Sea Research, 64(3): 213-223.
6 - Zhang, G., Zhang, J., Liu, S., Ren, J., Xu, J., & Zhang,
F., 2008. Methane in the Changjiang (Yangtze River)
Estuary and its adjacent marine area: riverine input,
sediment release and atmospheric fluxes. Biogeochemistry,
91(1): 71-84.
7 - Forster, G., Upstill-Goddard, R. C., Gist, N., Robinson,
C., Uher, G., & Woodward, E. M. S., 2009. Nitrous oxide
and methane in the Atlantic Ocean between 50 N and 52 S:
Latitudinal distribution and sea-to-air flux. Deep Sea
Research Part II: Topical Studies in Oceanography,
56(15): 964-976.
8 - Bange, H. W., Bartell, U. H., Rapsomanikis, S., &
Andreae, M. O., 1994. Methane in the Baltic and North
Seas and a reassessment of the marine emissions of
methane. Global Biogeochemical Cycles, 8(4): 465-480.
59
(2016)
¿Es el análisis de agua adecuado para el estudio de contaminantes orgánicos
en áreas costeras como el Mar Menor?
Víctor M. León1, Rubén Moreno-González1 & Juan A. Campillo1
1
Instituto Español de Oceanografía, Centro Oceanográfico de Murcia, C/ Varadero 1, San Pedro del Pinatar, 30740 Murcia,
Spain.
RESUMEN
En este trabajo se analiza la presencia y distribución de contaminantes orgánicos regulados y de interés emergente
en agua de mar del Mar Menor y se evalua su variabilidad diaria y estacional en puntos con distinto grado de
exposición a los principales focos de contaminación. Este estudio revela la gran variabilidad temporal de la
concentración de contaminantes orgánicos en esta matriz (pesticidas, PAHs, fármacos, etc.). Por tanto se evidencia
que un análisis puntual de agua de mar no es representativo para la evaluación de contaminantes orgánicos
hidrofóbicos en el medio, siendo necesario un muestreo intensivo o bien el uso de muestras que integren la
contaminación en un periodo de tiempo (sedimento, biota o muestreadores pasivos). Estos resultados evidencian
la necesidad de desarrollar criterios ambientales de referencia para estas matrices, más que para la matriz agua
como se ha hecho hasta ahora a nivel europeo en la Directiva Marco de Agua. También se han detectado
variaciones estacionales para estos compuestos en agua de mar, sedimento y biota, evidenciando la capacidad de
las altas temperaturas y la irradiación solar en el Mar Menor de reducir la carga de algunos contaminantes en
verano seguramente por procesos de degradación y/o volatilización. Además, se han observado diferentes patrones
de bioacumulación de algunos contaminantes orgánicos en moluscos y peces debido a factores ambientales,
fisiológicos y de comportamiento, lo que hace necesario identificar las especies más adecuadas para el
seguimiento de cada grupo de contaminantes en zonas costeras.
INTRODUCCIÓN
Las actividades humanas se concentran en las áreas
costeras que reciben descargas directas e indirectas de
contaminantes orgánicos, incluyendo hidrocarburos
aromáticos policíclicos, pesticidas, fármacos, tensioactivos,
productos de cuidado e higiene personal, etc. La evaluación
del estado del medio marino requiere del análisis de la
presencia, distribución y efectos de estos contaminantes,
para lo que es necesario disponer de muestras
representativas del medio. Muchos de estos contaminantes
tienden a adsorberse sobre el material particulado con el
que entran en contacto que tiende a depositarse en el
sedimento o bien entrar en la cadena trófica. Por ello las
matrices recomendadas por los convenios internacionales
(OSPAR, MED POL, etc.) para el seguimiento de los
contaminantes tradicionales como PAHs, contaminantes
organoclorados y metales traza incluyen el sedimento y la
biota. Sin embargo, en la legislación europea la matriz
preferente de referencia es el agua en ámbito continental y
marino (Directiva Marco de Agua), aunque la mayoría de
los contaminantes incluidos en la legislación tienen un
comportamiento eminentemente hidrofóbico y tiene una
baja solubilidad. Por ello es necesario poner en evidencia
qué información ofrece el análisis de una muestra de agua
y
cuáles son las alternativas que existen y que
comúnmente aplican los expertos en la materia.
Las presiones e impactos de las actividades humanas son
especialmente relevantes en áreas someras con una
capacidad limitada de dilución como es el caso de las
lagunas costeras. Estos sistemas presentan además
variaciones estacionales de las condiciones fisicoquímicas
y actividades humanas significativas, que deben ser
consideradas cuando se pretende evaluar su impacto en el
medio marino a través de programas de seguimiento de la
contaminación química.
En este estudio se caracteriza la variabilidad diaria y
estacional de contaminantes orgánicos regulados y de
interés emergente en agua de mar en puntos de muestreo
sometidos a distintos grados de contaminación química.
Los grupos de contaminantes orgánicos incluidos en este
trabajo han sido pesticidas organoclorados y de uso actual,
PAHs, bifenilos policlorados y fármacos, analizándose en
total más de 150 analitos. También se ha estudiado su
distribución estacional en agua, sedimento y biota del Mar
Menor con el objetivo de identificar la matriz o matrices
más adecuadas para cada grupo de contaminantes
estudiados.
MATERIAL Y MÉTODOS
Se han tomado muestras de agua de mar en 32 puntos cada
tres meses durante dos años (Figura 1) y en 19 puntos de
ellos también de sedimentos con periodicidad semestral.
En algunos puntos del principal aporte superficial (La
rambla del Albujón) y en la propia laguna se realizó un
muestreo diario de agua, para evaluar la representatividad
de una muestra puntual.
60
(2016)
200
Cage S3: 0.5 km El Albujón w.
Chlorpyrifos
Terbuthylazine
Concentration (ng L-1)
Además, se tomaron muestras de biota en primavera y
otoño de 2010 en 9 áreas de distintos puntos de la laguna,
incluyendo varias especies de moluscos y peces. Por último
se trasplantaron almejas a distintos puntos de la laguna para
evaluar la bioacumulación de contaminantes y los efectos
biológicos que estos puedan ocasionarles.
Propyzamide
150
Tributhylphosphate
100
50
0
0
Agua
Sedimento+
agua
Fig. 1. Distribución de los puntos de muestreo.
El análisis de diferentes grupos de contaminantes orgánicos
y matrices ha requerido de la aplicación de procedimientos
específicos según el caso. El análisis de PAHs, compuestos
organoclorados y pesticidas de uso actual en agua
superficial y agua de mar se ha realizado mediante
extracción con barras magnéticas recubiertas de
polidimetilsiloxano que son desorbidas térmicamente sobre
un cromatógrafo de gases con detección mediante
espectrometría de masas [1,2]. Sin embargo en el caso de
las muestras de fármacos se aplicó la cromatografía líquida
de ultra alta resolución y detección de espectrometría de
masas utilizando un pretratamiento específico para agua,
sedimento [3] y biota [4]. Por último, los PAHs y
compuestos organoclorados se extrajeron del sedimento y
biota mediante extracción Soxhlet, posterior purificación
por extracción en fase sólida y análisis por técnicas
cromatográficas específicas [5,6].
RESULTADOS Y DISCUSIÓN
La variación diaria y semanal es muy acusada para los
contaminantes orgánicos considerados en este estudio,
tanto en la desembocadura de la rambla del Albujón como
en la propia laguna. Como ejemplo de esta variabilidad se
muestra en la Fig. 2 la evolución de la concentración de
varios plaguicidas y el tributilfosfato. Estos resultados
evidencian que una muestra puntual de agua no es
representativa del medio para contaminantes con cierto
carácter hidrofóbico, siendo necesaria una serie temporal
50
100
Time (h)
150
200
más amplia para obtener datos ambientalmente relevantes
que nos den una visión de lo que ocurre realmente en esta
matriz.
Fig. 2. Variación de concentraciones de contaminantes en
agua en un punto próximo a la desembocadura de la rambla
del Albujón.
Así como alternativa se deben utilizar muestras
integradoras de la contaminación, como la biota, el
sedimento o los muestreadores pasivos que ofrecen un
valor integrador de la carga contaminante a la que han
estado expuestos. Se han observado diferentes patrones de
bioacumulación de algunos contaminantes orgánicos en
moluscos y peces debido a factores ambientales,
fisiolóficos y de comportamiento, siendo necesario
identificar y utilizar las especies más adecuadas para el
seguimiento de cada grupo de contaminantes en zonas
costeras. Por ello, se confirma la necesidad de seleccionar
las especies más adecuadas para el seguimiento de la
contaminación de acuerdo con las características de cada
grupo de contaminantes (especialmente los de interés
emergente).
Por tanto es fundamental la revisión de los indicadores y
criterios ambientales que proponen las directivas europeas
para la evaluación de la calidad del medio marino, ya que
básicamente proponen concentraciones de referencia en
agua, siendo necesarios valores específicos para matrices
ambientales más adecuadas como el sedimento o la biota.
AGRADECIMIENTOS
This work was supported by the Spanish Inter-Ministerial
Science and Technology Commission through the
‘IMPACTA’ project (CICYT, CTM2013-48194-C3-1-R)
and the DECOMAR project (CTM2008-01832), the Seneca
Foundation (Region of Murcia, Spain) through the
‘BIOMARO’ project (15398/PI/10), and by the European
Union through the European Regional Development Fund
(ERDF).
REFERENCIAS
1 - Moreno-González R, Campillo JA, García V & León, VM, 2013.
Seasonal input of regulated and emerging organic pollutants through surface
watercourses to a Mediterranean coastal lagoon. Chemosphere, 92: 247-257.
2 - Moreno-González, R., Campillo, J.A., León, V.M., 2013. Influence of an
intensive agricultural drainage basin on the seasonal of organic pollutants in
seawater from a Mediterranean coastal lagoon (Mar Menor, SE Spain). Mar.
Pollut. Bullet,. 77: 400–411.
3 - Gros M, Rodríguez-Mozaz S, Barceló D, 2012. Fast and comprehensive
multi-residue analysis of a broad range of human and veterinary
pharmaceuticals and some of their metabolites in surface and treated waters
by ultra-high-performance liquid chromatography coupled to quadrupolelinear ion trap tandem mass spectrometry. J Chromatogr. A, 22: 1-33.
4 - Huerta B, Jakimska A, Gros M, Rodriguez-Mozaz S & Barcelo D, 2013
Analysis of multi-class pharmaceuticals in fish tissues by ultra-highperformance liquid chromatography tandem mass spectrometry. J.
Chromatogr. A 1288: 63- 72.
61
(2016)
5 - Fernández B, Campillo JA, Martínez-Gómez C, Benedicto J, 2010.
Antioxidant responses in gills of mussel (Mytilus galloprovincialis) as
biomarkers of environmental stress along the Spanish Mediterranean coast.
Aquat. Toxicol., 99: 186–197.
6 - León VM, García I, Martínez-Gómez C, Campillo JA, Benedicto ,. 2014.
Heterogeneous distribution of polycyclic aromatic hydrocarbons in surface
sediments and red mullet along the Spanish Mediterranean coast. Mar Pollut
Bull
87:352-363.
62
(2016)
The vertical distribution of dissolved platinum in the West Atlantic Ocean:
evidence for a non-conservative behaviour
Daniel E. López-Sánchez1, Antonio Cobelo-García2
1
2
Universidad de Cádiz
Instituto de Investigacións Mariñas de Vigo (IIM-CSIC)
ABSTRACT
Dissolved Pt was analyzed in three depth profiles from the Western Atlantic Ocean (WAO) in the context
the Meridional Overturning Circulation (MOC). Samples were collected during the Dutch GAO2 GEOTRACES
cruise onboard the R/V Pelagia in 2011. One of the stations was located in the North Atlantic -39.399°W 47.801°
N, the second in the Central Atlantic -40.8835° W, 7.7664° N and the last in the South Atlantic -39.4425° W, 35.00835° S. We found important contribution of the water mass system over our profiles, the north profiles is
influenced mostly by the North Atlantic Depth Water (NADW), the second profile had influenced by the Antarctic
Intermediate Water (AAIW) and upper Circumpolar Deep Water (uCDW) and North Atlantic Deep Water
(NADW). Finally the third profile is affected similarly by the same water mass as the second profile, but
additionally we found the presence of the Antarctic Bottom Water (AABW). Dissolved Pt concentrations
observed in this study ranged from 0.11 to 0.37 pM, with an average value of 0.26 ± 0.06 pM (mean ± 1 SD;
n=59). Depth profiles showed a non-conservative behavior, in contrast with the only previous studies in the
Atlantic waters which were reported more than two decades ago. Possible causes for these discrepancies are
discussed.
INTRODUCTIÓN
Platinum is a highly siderophile metal and, as such, is one
of the least abundant elements at the Earth’s surface with a
typical crustal abundance of 0.5 ng g-1 [1]. In natural
waters, dissolved Pt typically displays picomolar and subpicomolar concentrations [1]. Current interest on the
investigation of the environmental Pt geochemistry relies
on the fact that its cycle at the Earth’s surface is greatly
impacted by anthropogenic activities, amounting up to, at
least, 80% of its total mobilization. Among these activities,
the use of Pt in automotive catalytic converters has been
identified as the major source of anthropogenic Pt released
into the environment; accordingly, elevated Pt
concentrations have been extensively reported in areas
close to vehicular traffic [1] but also evidence for a global
Pt environmental disturbance has been given [2].
Oceanic profiles of dissolved Pt have been reported for the
Pacific [3], Indian [4] and Atlantic [5] oceans, with
concentrations ranging from 0.2 to 1.6 pM. In these studies,
concentrations invariant with depth in the North Atlantic
[5] were reported; a scavenged-type profile in the Indian
Ocean [4] was observed, whereas for the Pacific the
available studies show discrepant behavior: recycled-type
[3] and conservative [6]. The behavior of platinum derived
from these studies is not oceanographically consistent with
respect to their basin-to-basin variation, suggesting the
possibility of error in some of the data.
In order to shed further light on the oceanic behaviour of
Pt, three vertical profiles were analysed in the West
Atlantic waters within the framework of the international
GEOTRACES program.
MATERIAL AND METHODS
The three profiles for dissolved Platinum (PtD) covered the
West Atlantic Ocean. Samples were collected during the
GEOTRACES GA02 transect from Iceland to Punta
Arenas (Chile). Samples were taken using 24 trace-metal
clean 24-L PVDF bottles mounted on a titanium frame with
a SEABIRD 911 CTD system and deployed with a Kevlar
hydrowire [7]. The complete ‘‘ultraclean CTD’’ was
immediately placed in an ISO Class 6 clean room
container, where samples for dissolved metals were filtered
directly from the PVDF samplers using 0.2 µm cartridges.
Samples were acidified to pH 1 using ultrapure HCl
(Seastar Chemicals). All sample processing for Pt
determination was carried out in a laminar flow bench
(ISO-5) housed inside an ISO-7 lab. Dissolved Pt was
analyzed by means of catalytic adsorptive cathodic
stripping voltammetry (Cat-AdCSV), using the procedure
described in [1]. Briefly, acidified (pH 1; HCl, Merck
Suprapur®) samples were UV-digested in quartz tubes
with PTFE caps using a 125-W high-pressure mercury
lamp for 2 hours. After digestion, sulphuric acid (final
concentration 0.5 M; Trace Select, Fluka), formaldehyde
(final concentration 3.5 mM; Riedel-de-Haen) and
hydrazine sulfate (final concentration 0.45 mM; Fluka)
were added. Samples were then transferred to a PTFE
voltammetric cell and a deposition potential of − 0.3 V (vs.
Ag/AgCl) was applied for 10–20 min depending on
platinum concentrations. After a quiescence of 10 s, the
63
(2016)
RESULTS AND DISCUSSION
Dissolved Pt concentrations observed in this study ranged
from 0.11 to 0.37 pM, with an average value of 0.26 ± 0.06
pM (mean ± 1 SD; n=59). Concentrations at stations Leg1
and Leg2 (Figure 1) were almost identical, with values of
0.23 ± 0.05 pM (0.12-0.30 pM; n=18) and 0.22 ± 0.05 pM
(0.11-0.32 pM; n=17) respectively, whereas for Leg3
values were significantly (two-tailed P<0.0001) higher:
0.31 ± 0.03 pM (0.25-0.37 pM; n=24). These values are in
close agreement with the only dataset to date in the
Atlantic Ocean [5], which reported concentrations of 0.26 ±
0.08 pM (0.14-0.39 pM; n=19) in the water column near
Bermuda (32°10’N, 64°30’W) and 0.30 ± 0.07 pM (0.110.41 pM; n=17) near Azores (26°20’N, 33°40’W).
Recently an oceanic residence time for Pt of 2.4 ± 1.0 104
years was estimated [8], i.e. lower than the long residence
times (>105 years) typical of conservative elements but
within the typical range of recycled elements (103-105
years), and higher than for scavenged elements (<103
years). This cast some doubt on the previous results
reporting conservative or scavenged-type behaviour, but is
in agreement with the results found in our study. However,
more studies are needed in order to better understand the
factors controlling the oceanic behaviour of Pt.
0
0.0
0.1
Pt (pM)
0.2
0.3
0.4
0.5
1000
Depth (m)
potential was scanned to − 1.1 V in the differential pulse
mode and the Pt peak at about − 0.90 V quantified. A
Metrohm 663 VA polarographic stand (Herisau,
Switzerland), equipped with a HMDE (working electrode),
a Ag/AgCl (reference electrode) and a glassy carbon rod
(auxiliary electrode) was used. The detection limit of the
technique, expressed as three times the standard deviation
of the blanks, was 0.02 pM. Given that there is no water
reference material for Pt at present, the accuracy of the
analytical procedure was checked by means of the analysis
of spiked oceanic waters, obtaining typical recoveries of
102 ± 10 %.
2000
3000
4000
5000
Fig. 2. Vertical profile of dissolved Pt at Leg2 station (Fig.
1), where a typical recycled-type distribution is observed.
ACKNOWLEDGEMENTS
We would like to thank the CSIC for the PhD grant to
DELS, and M. Rijkenberg (NIOZ) and the Dutch
GEOTRACES program for providing the samples.
Fig. 1. Location of the stations (red broken lines) in the
West Atlantic Ocean where dissolved Pt was analyzed.
However, in our study we found a non-conservative
behaviour for Pt in the West Atlantic Ocean. Figure 2
shows the profile for the Central Atlantic Station (Leg2),
indicating a recycled-type behaviour which is in contrast
with the conservative behaviour reported in the only
previous dataset for Pt in the Atlantic more than two
decades ago [5]. Accordingly, the previous studies on the
oceanic behavior of Pt do not show a consistent pattern,
with conservative, scavenged-type and recycled-type
reported [3-6] for the Indian, Pacific and Atlantic waters.
REFERENCES
1. Cobelo-García, A., et al. Mar Chem, 2013. 150: 11
2. Soyol-Erdene, T.-O., et al. Environ Sci Technol, 2011.
45: 5929
3. Goldberg, E.D., et al. App Geochem, 1986. 1: 227.
4. van den Berg, C.M. and G.S. Jacinto. Anal chim acta,
1988. 211: 129
5. Colodner, D.C., et al. Anal Chem, 1993. 65: 1419
6. Suzuki, A., et al. Mar Chem, 2014. 166: 114
7. Rijkenberg, M.J.A., et al. Mar Chem, 2015. 177: 501
8. Soyol-Erdene, T.-O., Huh, Y. Geochem Geophys
Geosyst, 2012. 13: Q06009
64
(2016)
Assessing the reactivity of DOM along the Levantine intermediate waters of
the Mediterranean Sea
Alba María Martínez-Pérez1, Mar Nieto-Cid1 & Xosé Antón Álvarez-Salgado1
1
Instituto de Investigaciones Marinas, CSIC
ABSTRACT
In spite of the key role played by marine dissolved organic matter (DOM) in the global carbon cycle, the
bioavailability of this pool in the dark ocean is still poorly understood. Current hypotheses, posed by the “sizereactivity continuum” and the “microbial carbon pump” conceptual frameworks, need to be tested experimentally.
In this context, the Mediterranean Sea —often considered as a laboratory basin to investigate processes occurring
at the World Ocean scale— has been chosen to test these hypotheses. The reactivity of DOM was followed at the
depth level of the Levantine Intermediate water (LIW) in their route across the entire Mediterranean Sea. We sizefractionated the DOM of the LIW using an efficient ultrafiltration cell (cut off 1000 Da) observing a significant
decrease of the percentage of high molecular weight DOM (HMW-DOM) with increasing apparent oxygen
utilization (AOU) rates. HMW-DOM was responsible for 29% of the oxygen consumption at the LIW level,
supporting the “size reactivity continuum” postulate.
INTRODUCTION
MATERIAL AND METHODS
Marine dissolved organic matter (DOM) is one of the
largest and least understood reservoirs of reduced carbon
on the Earth’s surface [1]. At 662 Pg C, DOM represents
96% of the total organic carbon in the oceans [2]. It is
produced mainly in the epipelagic layer (0–150 m depth) as
a result of phytoplankton photosynthesis and subsequent
food web interactions [3]. The aim of this work is to test
the “size-reactivity continuum” hypothesis [4]. According
to this postulate, changes in the bioavailability of DOM
could be explained by varying proportions of more labile
high molecular weight DOM (HMW-DOM) compared
with slower degrading low molecular weight DOM (LMWDOM).
We have used tangential flow ultrafiltration with a 1 kDa
membrane cut-off to separate the high and low molecular
weight fractions of the DOM in the Med Sea. The Med Sea
is a semi-enclose basin open to the Atlantic Ocean through
the Gibraltar Strait. It is composed of two basins of similar
size, the eastern and western basins separated by the Sicily
Strait. It is characterized by relatively high salinity and low
nutrient concentrations and it is used as a lab basin to test
processes happening at a global scale [5]. If the “sizereactivity continuum” hypothesis is correct, the Atlantic
water entering the Strait of Gibraltar should transport high
concentrations of DOM with a high average molecular
weight, which would be progressively consumed within the
overturning cell in such a way that the Levantine
Intermediate water (LIW) that leaves the Strait of Gibraltar
should transport low concentrations of DOM with low
average molecular weight. To test this hypothesis we
followed the LIW along the whole Med Sea.
Water samples were collected during the transMediterranean cruise HOTMIX aboard R/V Sarmiento de
Gamboa (Heraklion, Crete, 27 April 2014 – Las Palmas,
Canary Islands, 29 May 2014).
Fig. 1 Study area and sampling stations. Black circles
depict the whole cruise stations and red asterisks represent
the stations for the DOM size-fractionation.
At each station, full-depth continuous conductivitytemperature-depth (SBE 911 plus CTD probe), dissolved
oxygen and chlorophyll fluorescence profiles were
recorded. These probes were attached to a rosette sampler
(SBE 38) equipped with 24 (12 litres) Niskin bottles. Water
samples were collected to analyse salinity (S), dissolved
oxygen (DO) and chlorophyll a (Chl a) and calibrate the
CTD sensors for conductivity, DO and fluorescence.
Samples for S were measured with a Guildline Portasal
salinometer Model 8410A. Dissolved oxygen was
determined following the Winkler potentiometric method
modified after [6]. The Apparent oxygen utilization (AOU)
was calculated as in [7]. For the determination of dissolved
organic carbon (DOC), and DOM fractionation, seawater
samples were collected in 5-litres acid-cleaned
65
(2016)
polycarbonate carboys and then filtered through
precombusted GF/F filters. Aliquots of 10 mL of the
filtrate were collected for DOC quantification in
precombusted glass ampoules. These samples were
acidified to pH < 2 and the ampoules were heat-sealed and
stored in the dark at 4 °C until analysis. DOC concentration
was determined with a Shimadzu TOC-V organic carbon
analyzer by high temperature catalytic oxidation (HTCO).
At nine stations samples were taken for the DOM sizefractionation. Two-liter aliquots of the filtrate were
collected in acid-cleaned Teflon bottles for DOM
fractionation using an ultrafiltration cell (Millipore, 2000)
equipped with a membrane of 1kDa cut-off (Millipore,
PLAC 150 mm) and applying 55 psi of pressure with N2.
RESULTS AND DISCUSSION
In this study we focused on the LIW as this is the unique
water mass found throughout the entire Med Sea and can
be easily traced by its characteristic S maximum between
200–500 m depth [8]. The DOC concentration in the core
of this water mass decreased from 60.2 ± 0.9 in the
easternmost station down to 45.7 ± 0.6 µM in the
Mediterranean water (MW) in the Atlantic Ocean. By
contrast, the AOU increased from 18.1 up to 93 µM. These
results indicate that DOM in the LIW is subject to
progressive microbial degradation as it displaced from the
Levantine Sea to the Strait of Gibraltar. To test which is the
DOM fraction that is more susceptible to microbial
degradation, we size-fractionated the DOM pool into
HMW (> 1kDa) and LMW (< 1kDa) fractions.
Fig. 2. AOU vs DOC for DOM fractionation
HMW-DOM represented 76% at the easternmost station
(stn 1) and 58% at the MW in the Atlantic Ocean (stn 25)
and was inversely correlated with the AOU. Fig. 2 shows
that the consumption of the DOC pool (green dots) occurs
in parallel to the increasing oxygen demand. Note that the
decline of the DOC pool correspond to the decay of the
HMW-DOM fraction (purple dots), whereas the LMWDOM fraction (yellow dots) did not show significant
decrease; even a minor, but not significant, increase with
AOU can be envisaged. The slopes of the relationships of
DOC and HMW-DOC with AOU (-0.21 mol C mol O2–1),
indicate that 29% of the organic matter remineralization is
due to DOM, and particularly to the HMW-DOM. This
result is in accordance with [9] in the Eastern Med Sea.
Moreover, this result supports the “size-reactivity
continuum” hypothesis. This outcome does not exclude the
microbial carbon pump hypothesis that assumes refractory
HMW-DOM production during remineralization processes.
ACKNOWLEDGEMENTS
The authors are grateful to the Captain, crew, technicians
and scientists aboard the R/V Sarmiento de Gamboa. We
thank M.J. Pazó and V. Vieitez for DOC measurements.
This work was financed by the project HOTMIX (grant
number CTM2011-30010-C02-02-MAR). A.M.M.-P. was
funded by a predoctoral fellowship and a short stay
fellowship from the Mineco. M.N.-C. was supported by the
CSIC Program “Junta para la Ampliación de Estudios” cofinanced by the ESF.
REFERENCES
1 - Hedges, J.I., 1992. Global biogeochemical cycles:
progress and problems. Mar. Chem. 39, 67-93.
2 - Hansell, D.A., Carlson, C.A., Repeta, D.J., Schlitzer,
R., 2009. Dissolved organic matter in the ocean a
controversy stimulates new insights. Oceanography 22,
202-211.
3 - Carlson, C.A., 2002. Chapter 4 - Production and
Removal Processes. In: Hansell, D. A., Carlson, C. A.
(Eds.), Biogeochemistry of Marine Dissolved Organic
Matter. Academic Press, San Diego, pp. 91-151.
4 - Amon, R.M.W., Benner, R., 1996. Bacterial utilization
of different size classes of dissolved organic matter.
Limnol. Oceanogr. 41, 41-51.
5 - Bergamasco, A., Malanotte-Rizzoli, P., 2010. The
circulation of the Mediterranean Sea: a historical review of
experimental investigations. Adv. Oceanogr. Limnol. 1, 1128.
6 - Langdon, C., 2010. Determination of dissolved oxygen
in seawater by Winkler titration using the amperometric
technique. GO-SHIP repeat hydrography manual: a
collection of expert reports and guidelines edited by BM
Sloyan and C. Sabine, IOC/IOCCP, Paris.
7 – Benson, B., Krause, J. 1984. The concentration and
isotopic fractionation of oxygen dissolved infreshwater and
seawater in equilibrium with the atmosphere. Limnol.
Oceanogr. 29 (3), 620-632.
8 - Roether, W., Klein, B., Beitzel, V., Manca, B.B. 1998.
Property distributions and transient-tracer ages in
Levantine Intermediate Water in the Eastern
Mediterranean. J. Marine Syst. 18, 71-87.
9 - Meador, T.B., Gogou, A., Spyres, G., Herndl, G.J.,
Krasakopoulou, E., Psarra, S., Yokokawa, T., De Corte, D.,
Zervakis, V., Repeta, D.J., 2010. Biogeochemical
relationships between ultrafiltered dissolved organic matter
and picoplankton activity in the Eastern Mediterranean
Sea. Deep-Sea Res. II Top. Stud. Oceanogr. 57, 1460-1477.
66
(2016)
Temporal evolution of Rare-earth Elements concentrations in the
southwestern Iberian Peninsula shelf: sources and distribution
Mário Mil-Homens1,2, Pedro Brito1,2, Filipa Naughton1, Teresa Drago1 Joana Raimundo1,2,
Carlos Vale2 & Miguel Caetano1,2
1
2
IPMA, Portuguese Institute of Sea and Atmosphere, Portugal
CIIMAR, Interdisciplinary Centre of Marine and Environmental Research, University of Porto, Portugal
ABSTRACT
Concentrations of rare-earth elements (REEs), Al, Fe and grain-size parameters were measured in a 5 m sediment
core collected in the shelf area offshore of the southwestern Iberian Peninsula. The core VBC2 was dated using a
combination of 210Pb and 14C determinations. The highest fine-grained contents together with Al, Fe and REE
concentrations were found towards the Present. This suggesting less dynamic environmental/oceanographic
conditions that favoured the depositional processes associated with the stabilization of shoreline at 3050BC and a
major detrital contributions derived from the shift to humid conditions “Little Ice Age” (LIA, 1350AD-1900AD)
at approximaltey 1350AD. The intense mining activities in the Iberian Pyrite Belt (IPB) sulfide massive deposits
(between the 1860s to the 1960s) mobilized great amount of enriched REE particles to the environment. The
NASC-normalized REE pattern is similar to those found in the Guadiana estuarine sediments pointing to a major
estuarine contribution to the sediment load deposited in the adjacent coastal zone.
INTRODUCTION
Rare-earth elements (REE) have been used to study trace
sediment provenance[1], due to their low mobility during
sedimentary processes and short residence times in
seawater [2]. The lower sector of the Guadiana watershed
crosses the IPB, one of the largest world's massive sulfide
province, mined since the Copper Age (3000 to 2000BC)
until the Present day[3].
In this study is reported the temporal distribution of grainsize, Al, Fe and REE concentrations during the last 9.5 kyrs
in a sediment core of the Algarve shelf area of the
southwesten Europe. The main objectives are to infer the
sediment provenance and the depositional conditions.
MATERIAL AND METHODS
The VC2B sediment vibrocore with five meters length was
retrieved in the shelf area offshore of southwest Iberia
Peninsula at 96 m water depth (36º53'16.24''N,
8º04'06.39''W). Grain-size analysis was performed using a
Malvern Mastersizer 2000 laser diffraction particle sizer
with a measuring range of 0.014 μm–2000 mm. The age
model was based on six 210Pb and 226Ra determinations in
the upper 30 cm and on eight accelerator mass
spectrometry 14C dates (AMSC14) on marine material
(shell
and
planktonic
foraminifera).
Elemental
concentrations were obtained after total digestion of 0.1 g
of grounded sample (< 2 mm) with a mixture Aqua Regia
and HF in closed Teflon bombs at 100 ºC for 1 hour.
Aluminium was determined by flame atomic absorption
spectrometry (Perkin Elmer AA100) using a nitrous oxide-
acetylene while Fe was measured with an air acetylene
flame. Concentrations of REE elements were measured in
the same sample solutions using a quadrupole ICP-MS
(Thermo Elemental, X-Series).
RESULTS AND DISCUSSION
Sediment samples are dominated by silt. Fine-grained
contents varying between 75% and 98% presenting an
increase towards the surface. Concentrations of Al were
relatively constant until c.a. 1350AD followed by an
increase towards the Present (max 7.8%). This variability
parallels the grain-size profile pointing to the enrichment of
fine-grained particles in the last 1400 years. Concentrations
of Fe present similar down-core trends to Al and significant
correlations (p< 0.01; r2=0.94). These correlations suggest
the association with fine-grained particles reflecting major
detrital contributions and/or a decrease of energetic
conditions related to the coastline evolution and to the
transition from dry to humid conditions. The coastline
stabilization at about 3050BC[4] is followed by a decrease
of the mean-grain size values from 3050BC to 1400AD.
Above this age, it is observed a significant increase of the
fine fraction due to the proximity to Guadiana river mouth
and to high river discharge associated with wetter climatic
conditions of the LIA.
The sum of REE (∑REE) in sediment core samples varied
between 69 and 150 mg kg-1. The ∑REE are lower than the
mean values for the Guadiana estuarine sediments (212 mg
kg-1[5] and 182 mg kg-1[6]). In average for all core depths,
L-REE (La - Pr series) corresponds to 69%, M-REE (Nd Dy series) to 28% and H-REE (Ho - Lu series) to 3% of the
67
(2016)
∑REE concentration. All REE are significantly correlated
among them with Al and Fe (0.89< r2 <1.00, p< 0.01).
In general, the REE depth variability is marked by an
increase towards the Present. The low concentrations are
found in bottom sediments with the highest sand contents,
suggesting that REE pattern is influenced by grain-size.
The significant relationships, the normalized profiles to Al
still present high variability, suggesting other sediment
component than fine-grained particles influencing the REE
distributions. Besides clays, heavy minerals (e.g.,
monazite, zircon) may act as important REE carriers [7].
The normalization of REE concentration to a shale
composite, North American Shale Composite (NASC), is
usually used to identify sedimentary patterns through an
enrichment or deficiency (fractionation process) of a single
element or group of elements[8]. A clear temporal
separation of sediment layers is found (fig. 1): (i) before
3500BC marked by deposition of coarse material; (ii)
between 3500BC and 1450AD evidencing grain-sizes
fluctuating around a mean value; (iii) between 1450AD and
1850AD sediment deposition shows high fine-grained
particles and (iv) younger than 1850AD associated with the
mobilization of great amount of particulated materials
released to the environment by modern mining activities.
These patterns were always lower than 1 and decreased
with depths, indicating no enrichment in REE
concentrations with respect to the shale and also increase
dilution by marine carbonate and sand contents. In spite of
the clear separation in four periods the L-REE is always
enriched relatively to H-REE. Sediments older than
3500BC showed an almost flat profile of the La - Lu series
indicating a similar proportion of L-REE to H-REE.
NASC-normalized REE pattern is similar to those found in
the Guadiana estuarine sediments [5,6] with the enrichment
of L-REE and M-REE relative H-REE.
Fig.
1.
NASC-normalized
Normalization values from [9].
REE
distributions.
The Ce anomalies (Ce/Ce*=3 x CeNASC/(2 x LaNASC +
NdNASC [10]) are closer to 1 but ratios shows a gradual
increasing trend from 1800 AD to the Present. Ce/Ce* is
positively and significantly correlated with the ∑REE
(r2=0.89) and with Al (r2=0.85) suggesting that the degree
of Ce depletion is relative to the decrease of REE
concentrations.
The
estimated
Eu
anomalies
(Eu/Eu*=EuNASC/(sqrt(LaNASC x PrNASC)) [11]) were also
closer to the unit (0.97-1.12) indicating a lack of significant
Eu anomalies. Nevertheless, values decrease after 1940 AD
pointing to an increased exchange and transport of this
element to the water column. Interestingly, Ce and Eu
anomalies are decoupled being the Ce/Ce* the
differentiator. This suggests that Eu/Eu* may occur in any
sediment layer while Ce/Ce* is characteristic of a certain
period of time.
ACKNOWLEDGMENTS
This study was financially supported by the Foundation for
Science and Technology (FCT) through the POPEI project
(FCT/PDCT/ MAR/55618/2004). CLIMHOL project
(PTDC/AAC-CLI/100157/2008) supported the financial
costs with AMSC14 measurements. This research was
partially supported
by the
Strategic
Funding
UID/Multi/04423/2013 through national funds provided by
FCT and European Regional Development Fund, in the
framework of the programme PT2020.
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D, Scott DB, Fernandes SG, 2002. Postglacial sea-level
rise and sedimentary response in the Guadiana Estuary,
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5 - Pérez-López R, Delgado J, Nieto JM, Márquez-García
B, 2010. Rare earth element geochemistry of sulphide
weathering in the São Domingos mine area (Iberian Pyrite
Belt): A proxy for fluid-rock interaction and ancient
mining pollution. Chem Geol, 276:29-40.
6 - Delgado J, Pérez-López R, Galván L, Nieto JM, Boski
T, 2012. Enrichment of rare earth elements as
environmental tracers of contamination by acid mine
drainage in salt marshes: A new perspective. Mar Poll Bull
64:1799-1808.
7 - Yang SY, Jung HS, Choi MS, Li CX, 2002. The rare
earth element compositions of the Changjiang (Yangtze)
and Huanghe (Yellow) river sediments. Earth Planet Sc
Lett, 201:407-419.
8 - Haskin LA, Haskin MA, Frey FA, Wildeman TR, 1968.
Relative and absolute terrestrial abundances of the rare
earths. In: Origin and Distribution of the Elements. Ahrens
LH (Ed) Pergamon press, New York:889–911.
9 - McLennan SM, 1989. Rare earth elements in
sedimentary rocks; influence of provenance and
sedimentary processes. Rev Mineral Geochem, 21:169-200.
10 - Elderfield H, Greaves MJ, 1982. The Rare-Earth
Elements in Sea-Water. Nature, 296:214-219.
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11 - Taylor S, McLennan S, 1985. The continental crust:
Its composition and evolution. Blackwell (Oxford),
London.312
69
(2016)
Improving UV-based technologies for marine water disinfection:
Application to Ballast Water
J. Moreno-Andrés, L. Romero-Martinez, A. Acevedo-Merino & E. Nebot
Department of Environmental Technologies, Faculty of Marine and Environmental Sciences, Campus de Excelencia
Internacional del Mar (CEI·MAR), University of Cádiz. Campus Universitario Puerto Real, 11510 Puerto Real, Cádiz, Spain.
ABSTRACT
Water on ships is employed in the majority of the vessel activities, being necessary to carry out a correct
management of it, which involves marine water treatment. From the point of view on the selection of
technologies to be used on board, they have to adapt to special requirements such as little space available,
operation simplicity, treatment effectiveness, etc. Among the mainly water streams generated on vessels, appears
the ballast water as emerging challenge (especially on cargo ships) due to the transport of invasive species, and
the large impact that the ballast water discharges could cause on ecosystem and human activities. To avoid this
problem, it must be implemented ballast water treatment before water discharge according with the nearly
coming into force Ballast Water Management Convention - BWMC. The effectiveness of microbiological
disinfection by UV-based treatments (Advanced Oxidation Processes –AOP) has been evaluated in this study.
The aim of it was to investigate inactivation rates of microbiological indicator E. faecalis, established on
BWMC, by means of photolytic (UV-C), photocatalytic (UV-C/TiO2), and photolysis (UV-C/H2O2). A
comparison between photo-chemical processes was assessed by using flow-through UV-reactor: working at
different flow conditions with the goal of treatment optimization. Both photolysis and photocatalysis were far
more effective, as it reduced dose requirements by up to 57.8% (UV-C/TiO2) and 30.5% (UV-C/H2O2). The
results highlight the potential application of environmentally-friendly, innovative, and advanced technologies for
marine water treatment.
INTRODUCTION
Maritime transport is continuously expanding through the
globe, it actually covers about 90% of total world
merchandise [1]. Besides, the cruise tourism industry, has
experienced a boom in recent years: the number of people
who have chosen to spend their holidays aboard a cruise
ship has multiplied by four in the last two decades [2],
meaning in exponential growth. The water on vessels is
used for almost all activities carry out on board, and it
implies the needs to discharge it. These environmental
pressures could be enough to constitute a health hazard to
the ecosystem and increase marine pollution.
Among the mainly water streams generated on vessels,
appears the ballast water as emerging challenge. Ballast
water is needed on oceangoing vessels to ensure ship
stability and buoyancy; when ballast water is released into
far ecosystems, the organisms included herein could find a
way that enables them to develop and spread into the new
habitat, becoming invasive [3]. Invasive aquatic species
involve a global challenge and one of the most severe
pollution problem facing the world´s oceans [3]: “one
single ballast tank was estimated to contain more than 300
million harmful organisms that could be developed into
confirmed toxic cultures” [4]. So, is essential to develop
ballast water treatments (BWTs) and management
strategies to minimize the spread of organisms in ballast
water.
Usually, BWTs consists in a filtration step followed by a
disinfection phase; it could be applied physically, chemical
or physical-chemically [5]. A BWTs must address a
number of parameters established on Ballast Water
Management Convention (BWMC), developed by IMO
and adopted in 2004 [6]: it established guidelines including
different indicators and discharge limits for ballast water.
According to these procedures, the goal of this study is to
optimize different UV-based treatments for ballast water
disinfection, in order to promote the application of
sustainable and advanced technologies (AOPs) for these
purposes.
The specific goal is to investigate inactivation rates of
microbiological indicator E. faecalis, established on
BWMC, by means of photolytic (UV-C), photocatalytic
(UV-C/TiO2), and photolysis (UV-C/H2O2).
MATERIAL AND METHODS
- Microbiological procedures: Pure cultures of E. faecalis
(ATCC 27285) were inoculated to Artificial Seawater for
reach concentrations around 106-107 CFU/mL. Posttreatment analysis was assessed by membrane filtration
technique.
70
(2016)
- UV-Reactor´s: Two tubular UV reactors were used on
experiments: A PVC reactor for UV and UV/H2O2
treatment and a photocatalytic reactor (with fixed TiO2)
for UV/TiO2 treatment. All of them are tubular reactors
operating in continuous-flow. The dose was calculated as
the product of mean intensity and Theoretical Retention
Time.
- Experimental: The experimental procedure consisted of
applying three different treatments, at different UV-C
doses. In the case of hydrogen peroxide, it was added in a
single dosage before the UV irradiation to reach a H2O2
concentration of 5 mg/L.
- Data treatment: The effectiveness of the treatment was
determined by logarithmic reduction of the survival of
microorganisms: log (N/N0). It was modeled with
GinaFiT tool [7].
RESULTS AND DISCUSSION
The results obtained from different treatments are shown in
Fig. 1, in which the logarithm inactivation versus the UV
dose received is represented.
Photocatalytic treatment shows even a better improvement
of UV light: 57.8% of reduction in UV dose for reach 4Log inactivation. The generation of ·OH on a fixed catalyst
is derived from light incidence, it have the advantage of no
chemicals added, but some studies reflect the loss
efficiency in marine waters because of catalyst aging [9].
Table 1. Kinetic and statistical parameters predicted by
fitting of disinfection experimental data.
Treatment
S.L. (mJ/cm2)
kmáx (cm2/mJ)
R2
± S.E.
± S.E.
UV
5.92 ± 2.94
0.40 ± 0.03
0.957
UV/H2O2
1.91 ± 2.14
0.51 ± 0.03
0.976
UV/TiO2
2.32 ± 1.54
0.93 ± 0.08
0.965
Both AOPs are effective in disinfection of marine waters,
even with other indicators [9, 10]. Future studies could
evaluate the effect of organic matter present in real
seawater, or the possible synergistic effect that could have
the implementation of a combined treatment:
UV/TiO2/H2O2.
The results make possible the bet on more sustainable
technologies due to the absence of chemicals and low
formation of disinfection by-products; besides they can
adapt to on-board requirements. Moreover, it could
increase efficiency of water management on vessels and
explore the possibility of reusing some of the treated
streams and minimizing harmful discharges.
ACKNOWLEDGEMENTS
The Spanish Ministry of Economy and Competitiveness by
Ref. CTM2014-52116-R has supported the work.
REFERENCES
Fig. 1. a) UV dose necessary to inactivate up to four
logarithmic units for UV (purple), UV+H2O2 (blue), and
UV+TiO2 (orange) treatment. b) Modeled disinfection
profiles.
The best fitting model for the three disinfection profiles
was Log-linear+shoulder [7]: a shoulder region appear at
low UV doses (Table 1), where little or no inactivation
occurs, followed by a transition to log-linear yield.
Parameters of the model are shown in table 1; kinetic
constant was increased according to UV-C < UV-C/H2O2 <
UV-C/TiO2 meaning in an improvement on UV treatment
for both AOPs.
UV/H2O2 treatment generate hydroxyl radicals (·OH) by
photolysis of H2O2, these radicals could be the main
disinfection route. The results do not show significant
interference by application on marine waters. The
concentration used (5 ppm) was optimized in previous
studies in order to avoid ·OH recombination process and
high concentration of chemical [8]. The improvement was
of 30.5% of reduction in UV dose, for reach 4-Log
inactivation.
1 - IMO, 2012. International shipping facts and figures –
information resources on trade, safety, security,
environment.
2 - Cruise Market Watch, 2014. Growth of the Cruise Line
Industry. In: http://www.cruisemarketwatch.com/growth/.
3 - Werschkun B, Banerji S, Basurko OC et al., 2014.
Emerging risks from ballast water treatment: The run-up to
the International Ballast Water Management Convention.
Chemosphere 112:256–266.
4 - Belkin S & Colwell RR, 2005. Oceans and Health:
Pathogens in the Marine Environment. Springer
5 - Tsolaki E, Diamadopoulos E, 2010. Technologies for
ballast water treatment: a review. J Chem Technol
Biotechnol 85:19–32.
6 - IMO, 2004. International Convention for the Control
and Management of Ships’ Ballast Water and Sediments.
BWM/CONF/36.
7 - Geeraerd AH, Valdramidis VP, Van Impe JF, 2005.
GInaFiT, a freeware tool to assess non-log-linear microbial
survivor curves. Int J Food Microbiol 102:95–105.
8 - Moreno-Andrés J, Romero-Martínez L, AcevedoMerino A, Nebot E, 2016. Determining disinfection
efficiency on E. faecalis in saltwater by photolysis of
H2O2: Implications for ballast water treatment. Chem Eng
J 283:1339–1348.
9 - Rubio D, Casanueva JF, Nebot E, 2013. Improving UV
seawater disinfection with immobilized TiO2: Study of the
viability of photocatalysis (UV254/TiO2) as seawater
disinfection technology. J Photochem Photobiol A Chem
71
(2016)
271:16–23.
10 - Romero-Martínez L, Moreno-Andrés J, AcevedoMerino A, Nebot E, 2014. Improvement of ballast water
disinfection using a photocatalytic (UV-C + TiO 2 ) flowthrough reactor for saltwater treatment. J Chem Technol
Biotechnol
89:1203–1210.
72
(2016)
Identificación, cuantificación y análisis de ácidos grasos de la comunidad
fitoplanctónica del área de influencia del Guadalquivir en el Golfo de Cádiz
Identification, quantification and fatty acids analysis of the phytoplankton community of
the Guadalquivir influence area in the Gulf of Cadiz.
R. Muñoz-Lechuga (1), S. van Bergeijk (1), C. Vilas (1), R. Sánchez-Leal (2), C. Pérez-Gavilan
(1)
, J.P. Cañavate (1).
(4)
IFAPA Centro El Toruño, Junta de Andalucía. [email protected]
(5)
Instituto Español de Oceanografía – Centro Oceanográfico de Cádiz.
ABSTRACT
The phytoplankton community at the influence area of the Guadalquivir river is poorly studied in terms of genera
or species composition, or biomass and their role as major primary producers in the food web of the Gulf of
Cadiz has not been investigated. Despite its relevance, no work on species identification and quantification has
been carried out. In the present study, phytoplankton was sampled at 10 sites covering the entire area of
influence and the mouth of the estuary, from 10-80 m depth (surface and bottom), in the months of June, August
and November of 2013, using Niskin bottles collected with CTD casts. In total, 40 samples of phytoplankton
were analyzed, identifying genera or species (> 20 µm) and quantifying their abundance and biovolumes. At the
same time, fatty acid profiles of seston samples were analyzed to determine their role as the base of the food
web; these profiles are also explored as possible indicators of the phytoplankton community species
composition. Up to 40 genera or species were identified, mostly diatoms and dinoflagellates. Some of the
identified species are considered toxic to fish and bivalves showing significant values of abundances, as
Dinophysis spp., Pseudonitzchia spp. or Alexandrium spp. Spatial and seasonal gradients were observed in
phytoplankton distribution. Dinoflagellates were dominant in surface waters, near the coast, in June and August,
while diatoms were dominant in bottom waters, offshore and in November. Fatty acid markers of dinoflagellates
and diatoms confirm that distribution.
Key words: microfitoplancton, Golfo de Cádiz, distribución, cuantificación, ácidos grasos
INTRODUCCIÓN
El fitoplancton contribuye en gran medida en la producción
primaria de las bases de las redes tróficas marinas Pomeroy
(1974). Sin embargo, la comunidad fitoplanctónica en la
zona de influencia del rio Guadalquivir y el Golfo de Cádiz
está escasamente estudiada. El objetivo de este trabajo fue
estudiar la diversidad de géneros o especies, cuantificar la
biomasa y comprender su papel como productores
primarios en la red trófica del Golfo de Cádiz.
MATERIAL Y MÉTODOS
Las muestras fueron recogidas en la parte externa de la
zona de influencia del Golfo de Cádiz con el estuario del
Guadalquivir durante las campañas oceanográficas de
Junio, Agosto y Noviembre de 2013 del IEO-Cádiz durante
2013 (STOCA/INGRES, ECOCADIZ y ARSA en el marco
del proyecto ECOBOGUE (P. EXC. RNM 7467). Se
identificaron y cuantificaron géneros y especies, se
calcularon sus biovolúmenes, biomasa de carbono y se
realizaron análisis de ácidos grasos de muestras análogas.
Para la captura de imágenes se usó una cámara Leica
DFC420 incorporada al microscopio Leica DM5500B. El
método utilizado para la cuantificación de géneros y
especies fue el clásico de Utermöhl (1958), usando cámaras
de sedimentación. El cálculo del biovolumen individual de
cada especie/género se realizó asignando una figura
geométrica simple semejante a la forma del individuo
midiendo así sus dimensiones Hillebrand et al. (1999). Para
el cálculo de la biomasa de carbono se utilizó un ratio que
relaciona el biovolumen del individuo con su contenido en
C, para diatomáceos (0.15pg C μm−3) y no diatomáceos
(0.225 pg C μm-3) Reynolds (1984). Durante los muestreos
el CTD registró datos de temperatura, salinidad, oxígeno
disuelto y concentración de chl a en el agua. Además se
cogieron muestras de las botellas Niskin que se filtraron
por filtros GF/F (Whatman) calcinados para estimar los
sólidos en suspensión (filtros) y la concentración de NO3-,
NO2-, NH4+, PO43- y SiO32-, (filtrado). Para el análisis de
los ácidos grasos se recogieron muestras de seston en
botellas Niskin que se filtraron por filtros GF/F (Whatman)
calcinados y fueron analizados por cromatografía de gases.
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(2016)
El tratamiento estadístico ha sido realizado mediante los
softwares Microsoft Excel 2013, PRIMER-E y el software
libre R v3.0.1.
RESULTADOS Y DISCUSIÓN
Se identificaron más de 40 géneros o especies, la mayoría
de ellos dinoflagelados y diatomeas (Fig. 1). Los
dinoflagelados se encontraron entre el grupo de especies de
mayor tamaño. Entre las especies de menor tamaño se
identificó una clorofita y otra criptofita. Además se
localizaron
dos
especies
de
silicoflagelados
(Dictyochophyceae) y un flagelado heterótrofo, también
con esqueleto de silice (Ebria) de tamaños medios.
contribuyeron con más biomasa. Las altas densidades de
Gymnodinium spp. en junio, Alexandrium spp. y
Pseudonitzchia spp. en agosto (Fig. 1) y de Skeletonema
spp. en noviembre, parecen indicar “blooms” de esas
especies. Otras como Ceratium spp., Diploneis spp. o
Pleurosigma spp. fueron constantes en esos tres meses.
Analizando la distribución de las especies, se observan
diferencias espaciales y estacionales. El análisis
PERMANOVA nos indica una diferencia significativa
entre los meses de junio y noviembre con un (P=0.001)
(Fig. 2). La temperatura y las concentraciones de nitrato y
silicato tuvieron más relevancia en junio, sin embargo, en
los de noviembre fue el O2 disuelto y la concentración de
fosfato Espacialmente se observaron
diferencias entre
superficie y fondo en los tres meses, dominando los
dinoflagelados a poca profundidad y
Fig. 2. Análisis de componentes principales de los puntos
de muestreo y las variables ambientales de los meses de
junio y noviembre.
las diatomeas en el fondo. El análisis de ácidos grasos
confirma la presencia de marcadores para dinoflagelados
como el 18:5n3; 22:6n3; w3 y w6. Estos marcadores fueron
dominantes en el mes de junio para las radiales GD, PD y
MT. También se encontraron marcadores de diatomeas,
bacterias y material terrígeno como el 22:0; 18:0; BAME y
PU16. Presentaron dominancia en el mes de noviembre,
pero sólo en los radiales MT y PD.
Fig. 1. Biovolúmenes medios (μm3) por individuo de los
géneros y especies estudiados en este trabajo. Medias con
desviaciones y rangos. Los círculos representan valores
atípicos (Izq). Biomasa total media (ng de C L-1) frente a la
densidad total media (nº células L-1) del conjunto de
muestras del mes de agosto (Drch).
Closteriopsis sp. no se ha descrito para aguas marinas. Se
encontraron varias especies potencialmente tóxicas como
Alexandrium spp., Dinophysis spp. o Pseudonitzchia spp.
(Fig. 1). Varias diatomeas, como Pseudonitzchia spp. o
Skeletonema spp., alcanzaron densidades relativamente
altas, pero en términos de biomasa tuvieron poco peso en la
comunidad dado su pequeño tamaño. Por el contrario,
especies de mayor tamaño y menor abundancia
REFERENCIAS
Hillebrand, H. et al., 1999. Biovolume calculation for pelagic
and benthic microalgae. Journal of phycology, 35(2),
pp.403–424.
Pomeroy, L.R., 1974. The ocean’s food web, a changing
paradigm. Bioscience, 24(9), pp.499–504.
Reynolds, C.S., 1984. The ecology of freshwater
phytoplankton, Cambridge University Press.
Utermöhl, H., 1958. Zur vervollkommnung der quantitativen
phytoplankton-methodik. Mitteilung Internationale
Vereinigung fuer Theoretische unde Amgewandte
Limnologie, 9, pp.1–38.
74
(2016)
Grado de saturación del CaCO3 en el Golfo de Cádiz: una primera
aproximación
T. Ortega, A. Sierra, D. Jiménez-López, R. Ponce, A. Gómez-Parra, J. Forja
Dpto. Química-Física. CACYTMAR.Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz, Campus
Universitario Río San Pedro, 11510 – Puerto Real, Cádiz, Andalucía, España.
*Correo del autor: [email protected]
RESUMEN
En este trabajo se presenta un primer estudio sobre la distribución del grado de saturación del CaCO3 en el Golfo de Cádiz
para una situación de primavera. Se han tomado muestras a diferentes profundidades en la columna de agua en cinco
transectos, en los que se ha medido pH, alcalinidad total (AT) y la concentración de Ca2+. AT y la concentración de carbono
inorgánico disuelto (CID) varían entre 2316 y 2585 µmol kg-1, y 2134 y 2379 µmol kg-1 respectivamente. Se ha observado un
aumento de las concentraciones de AT y CID con la profundidad en la columna de agua, provocado por las propias
características hidrodinámicas de la zona y por la remineralización bentónica. Salvo en el transecto de Trafalgar, en el resto se
aprecia un aumento de la AT y el CID con la distancia a costa. La concentración de calcio varía entre 10.74 y 11.71 mmol kg-1
y aumenta con la distancia a costa en las aguas superficiales. Los mayores valores de Ca2+ se han medido en aguas profundas.
Las aguas superficiales en todos los transectos están sobresaturadas de CaCO3, con grados de saturación de calcita (Ωa) y
aragonito (Ωc) que varían entre 2.4 y 4.3, y 1.6 y 2.8 respectivamente. El comportamiento del grado de saturación del CaCO3
está condicionado por el pH, lo que provoca su aumento con la distancia a costa en las aguas superficiales y su disminución
con la profundidad.
INTRODUCCIÓN
La acidificación oceánica debida a la entrada de CO2
atmosférico está produciendo cambios en la química del
carbono del agua de mar, y sus efectos más inmediatos son
la disminución del pH y de la capacidad tampón del océano
debido a la pérdida de iones carbonato (CO2 + CO32- + H2O
= 2 HCO3-) [1, 2]. De hecho, esta acidificación puede estar
alterando las tasas de calcificación/disolución de algunos
organismos y del CaCO3 de los sedimentos, y se cree que
desempeñará un papel cada vez más importante en el ciclo
de CaCO3 oceánico de las próximas décadas [3].
Las aguas mediterráneas que entran en el Golfo de Cádiz a
través del Estrecho de Gibraltar están experimentando una
disminución anual de pH de -0.0044±0.00006 unidades,
que pueden interpretarse como un indicador del proceso de
acidificación de la cuenca [3], lo que puede repercutir de
una manera directa en el ciclo del CaCO3 de la zona. El
trabajo que se presenta, forma parte de un primer estudio
con el que se quiere conocer la influencia de la
acidificación oceánica y los aportes costeros sobre el grado
de saturación del CaCO3 en el Golfo de Cádiz.
El calcio es un elemento mayoritario del agua de mar, cuya
concentración apenas presenta cambios a nivel oceánico.
De ahí que en diferentes estudios su cuantificación se
realice a partir de la salinidad [1]. Sin embargo, en las
zonas costeras debido a los aportes terrestres y a procesos
de disolución/precipitación de CaCO3, este elemento puede
presentar considerables variaciones en su concentración.
De hecho, en la actualidad sólo existe un trabajo en el que
se han realizado medidas directas de calcio transoceánicas
[4] y en el que se ha obtenido que la profundidad de
saturación de la calcita y aragonito es menor si el calcio es
medido a si es calculado a partir de la salinidad. Como
parte de este trabajo, se ha optimizado la medida de calcio
en aguas costeras, lo que permitirá tener una mejor
cuantificación para trabajos futuros del grado de saturación
de CaCO3.
MATERIAL Y MÉTODOS
Fig. 1. Transectos de la zona de estudio del Golfo de Cádiz:
Guadiana (GU), Tinto y Odiel (TO), Guadalquivir (GD),
Sancti Petri (SP) y Trafalgar (TF).
En marzo de 2016 se ha realizado una primera campaña de
una serie de cuatro en el Golfo de Cádiz a bordo del B/O
Ángeles Alvariño. Se han considerado 5 secciones
75
(2016)
localizadas en la desembocadura del Tinto y Odiel (TO),
Guadiana (GU), Guadalquivir (GD), Sancti Petri (SP) y
Trafalgar (TF) (Figura 1). En cada una de ellas existe una
serie de estaciones fijas en la que se ha medido pH,
alcalinidad total y calcio en la columna de agua.
El pH y AT se analizaron mediante valoración
potenciométrica utilizando un valorador potenciométrico
(Metrohn 905) provisto de un electrodo combinado de
vidrio (Metrohm, ref 6.0210.100), calibrado previamente
en la escala total. Como valorante se utilizó HCl 0,1 M,
preparado en NaCl 0,7 M. Se tomaron muestras por
duplicado de 99,73 mL mediante una bureta. Las
concentraciones de carbono inorgánico disuelto y de
carbonato se obtuvieron a partir de la AT y pH utilizando
constantes disociación propuestas para la escala total [5].
La concentración de Ca2+ se midió por medio de una
valoración potenciométrica (Metrohn 905) utilizando un
electrodo selectivo de calcio (Metrohm, 6.0510.100). Se
utilizó como agente valorante ácido etilenglicol-bis-(2aminoetileter)-N,N,N´,N´ tetracético (EGTA) 0.01 M. A
las muestras se le añadió 10 mL de bórax 0.1 M para
tamponarlas a un pH próximo a 9. Loa análisis se han
realizado por duplicado utilizando aproximadamente 5 g
(±0,1 mg) de muestra por duplicado, y se ha obtenido una
desviación estándar media de ± 5 µM. El grado de
saturación de la calcita (Ωc) y aragonito (Ωa) se calculó a
partir de la concentración de calcio y carbonato y del
producto de solubilidad aparente del CaCO3 [6]. Se
consideró la variación de éste con la profundidad [7].
una clara disminución con la profundidad condicionada por
el descenso que experimenta el pH en la columna de agua.
En la Figura 2 se muestra a modo de ejemplo la variación
del Ωc en el transecto del Guadalquivir.
RESULTADOS Y DISCUSIÓN
Las concentraciones de AT y CID medidas en el Golfo de
Cádiz varían entre 2316 y 2585 µmol kg-1 y 2134 y 2379
µmol kg-1, respectivamente. Es en la sección de Trafalgar
donde se alcanzan los valores más elevados, posiblemente
debido a su mayor influencia del agua mediterránea. En los
cinco transectos estudiados se ha observado un aumento de
las concentraciones de AT y CID con la profundidad en la
columna de agua, provocado por la propia estratificación
de masas de agua en la zona y por la remineralización
bentónica. Salvo en el transecto de Trafalgar, en el resto se
aprecia un aumento de la AT y el CID con la distancia a
costa.
En la zona de estudio la concentración de Ca varía entre
10.74 y 11.71 mmol kg-1 y presenta un comportamiento
muy similar al de la salinidad en todos los transectos, con
un aumento de la concentración con la distancia a costa en
las aguas superficiales y los mayores valores en aguas
profundas.
El CO32- presenta un comportamiento similar al del pH,
con valores más elevados en aguas superficiales y en
aumento con la distancia a costa y disminución con la
profundidad en la columna de agua. Sus concentraciones
varían entre 104.2 y 176.0 µmol kg-1.
Las aguas superficiales del Golfo de Cádiz están
sobresaturadas de CaCO3, con valores de Ωa y Ωc que
varían entre 2.4 y 4.3, y 1.6 y 2.8 respectivamente. Los dos
grados de saturación presentan el mismo comportamiento,
REFERENCIAS
1 - Wanninkhof, R, Barbero, L, Byrne, R, Cai, W-J, Huang,
W-J, Zhang, J-Z, Baringer, L & Langdon, C, 2015.Ocean
acidification along the Gulf Coast and East Coast of the
USA. Continental Shelf Research, 98: 54–71.
2 - Flecha, S, Pérez, FF, García-Lafuente, J, Sammartino,
S, Ríos, AF & Huertas, IE, 2015. Trends of pH in the
Mediterranean Sea through high frequency observational
data: indication of ocean acidification in the basin.
Scientific Reports (DOI: 10.1038/srep16770)
3 - Smith, SV & Mackenzie. FT, 2015. The Role of CaCO3
Reactions in the Contemporary Oceanic CO2 Cycle. Aquat
Geochem. (DOI: 10.1007/s10498-015-9282-y)
4 - Rosón, G, Fernández-Guallart, E, Pérez, FF & Ríos, A
F, 2016. Calcium distribution in the subtropical Atlantic
Ocean: implications for calcium excess and saturation
horizons. Journal of Marine Systems, 158: 45-51.
5 - Lueker, TJ, Dickson, A.G. & Keeling, CD, 2000. Ocean
pCO2 calculated from dissolved inorganic carbon,
alkalinity, and equations for K1 and K2: validation based
on laboratory measurements of CO2 in gas and seawater at
equilibrium. Mar. Chem., 70:105-119.
6 - Mucci, A., 1983. The solubility of calcite and aragonite
in seawater at varius salinities, temperatures, and one
atmosphere total pressure. Am. J. Sci., 283:780-799.
7 - Zeebe, RE &Wolf-Gladrow, D, 2001.CO2 in Seawater:
Equilibrium, Kinetics, Isotopes. Elsevier Oceanography
Series 65, Amsterdam.
Profundidad (m)
Ωc
Distancia a costa (km)
Fig. 1. Variación del grado de saturación de la calcita (Ωc)
en la columna de agua en el transecto del Guadalquivir.
Los puntos negros muestran las estaciones de muestreo.
Ya que el grado de saturación se encuentra afectado en
gran medida por el pH, que a su vez presenta importantes
variaciones relacionadas con la actividad biológica en las
agua superficiales y se encuentra afectado por la entrada de
aguas continentales, se requiere un estudio estacional para
poder comprender la dinámica del CaCO3 en el Golfo de
Cádiz.
AGRADECIMIENTOS
El trabajo ha sido financiado por el Proyecto CICYT
CTM2014-59244-C3-1-R.
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Potencial de macroalgas como alimentos funcionales
Raquel Ortega1, Aroa López1, Adela Durá1, Argimiro Rivero1, Miguel Suárez de
Tangil1 & Juan Luis Gómez Pinchetti2
1
Grupo QUIMA-IOCAG, Universidad de Las Palmas de Gran Canaria, Campus de Tafira, 35017 Las Palmas de
Gran Canaria, Canary Islands, Spain
2
Banco Español de Algas, Instituto de Oceanografía y Cambio Global, Universidad de Las Palmas de Gran Canaria,
Muelle de Taliarte s/n, 35214 Telde, Canary Islands, Spain
RESUMEN
Los alimentos de origen vegetal despiertan un gran interés desde el punto de vista de la actividad
antioxidante. Enfermedades crónicas como el cáncer, alzheimer y enfermedades cardiovasculares entre
otras están causadas en gran parte por el estrés oxidativo y la peroxidación lipídica. Estos procesos se
inician, principalmente, por la aparición de radicales libres, que generan posteriormente Especies
Reactivas del Oxígeno (EROs) y producen un gran daño celular [1].
En este trabajo se plantea el estudio de la actividad antioxidante de algas macroscópicas con el fin de
comparar los resultados con los obtenidos para diferentes alimentos de origen vegetal: pimiento rojo
andaluz, la papaya, el pimiento verde, el pimiento amarillo, el maíz en lata y el maíz tierno maduro de
Canarias. Se utiliza la medida de la actividad antirradicalaria como un primer paso para determinar la
utilidad de estas macroalgas en la prevención de la oxidación molecular.
El método seleccionado para este estudio ha sido el descrito por Chu, Chang y Hsu [2], con ligeras
modificaciones. Se utiliza el radical 1,1-difenil-2-picrilhidrazilo (DPPH) como radical estable a
neutralizar por los antioxidantes de los extractos de las muestras. La disminución de la absorbancia
medida a 515 nm durante 10 minutos permite conocer la capacidad de los mencionados extractos de
inhibir el radical DPPH. En este ensayo se calcula el porcentaje de neutralización de radicales libres
(Radical Scavenging Activity, % RSA), y el tiempo que tarda en disminuir la concentración de DPPH en
un 50%, con el fin de obtener una visión de la rapidez y efectividad en la neutralización de los radicales
libres. Los valores de RSA fueron calculados en base a la siguiente ecuación:
RSA (%) = (1 – absorbancia de la muestra / absorbancia inicial de la disolución de DPPH) x 100.
ABSTRACT
Plant based foods have attracted a great interest because they exhibit antioxidant activity. Chronic
diseases such as cancer, Alzheimer's and cardiovascular diseases are largely caused by oxidative stress
and lipid peroxidation. These processes are mainly initiated by free radicals, which generate Reactive
Oxygen Species (ROS) and produce a cell damage [1].
In this study, macroalgae antioxidant activity was compared with those of extracts from different plant
materials: Andalusian red pepper, papaya, green pepper, yellow pepper, canned corn and sweet corn
mature from Canary Islands. The radical scavenging activity is used as a first step to determine the utility
of macroalgae in the prevention of molecular oxidation.
The method selected for this study was described by Chu, Chang and Hsu [2] and used with slight
modifications. 2,2-diphenyl-1-picrylhydrazyl (DPPH) is a stable free radical which has an unpaired
valence electron at one atom of nitrogen bridge. Scavenging of DPPH radical by the antioxidants from
extracts is the basis of this assay. The decrease in the absorbance was recorded at 515 nm for 10 min
giving the antioxidant activity of the extracts. The half-life (time required for reducing initial
concentration of DPPH by 50%) and the Radical Scavenging Activity, (% RSA) and were calculated in
order to gain an approach into the speed and effectiveness in neutralizing free radicals. RSA values were
calculated using the following equation:
RSA (%) = (1 – absorbance of sample/ absorbance of DPPH solution) x 100.
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REFERENCIAS
[1] Ferrazzano, G.F., Amato, I., Ingenito, A., Zarrelli,
A., Pinto, G. & Pollio, A. (2011). Plant
polyphenols and their anti-cariogenic properties:
a review. Molecules, 16, 1486-1507.
[2] Y.H. Chu, C.L. Chang, H.F. Hsu, Journal of the
Science of Food and Agriculture, 80 (2000) 561–
566.
78
(2016)
Copper effect on Fe(II) oxidation rate constant in seawater
Norma Pérez-Almeida1, Aridane G. González1, J. Magdalena Santana-Casiano1 and Melchor
González-Dávila1
1
Instituto de Oceanografía y Cambio Global (IOCAG). Universidad de Las Palmas de Gran Canaria. Las Palmas de Gran
Canaria, Spain
ABSTRACT
The interaction between the redox chemistry of Fe and Cu at nanomolar has been studied in UV-treated seawater.
The oxidation of Fe(II) was studied as a function of concentrations of Cu(II) and Cu(I) from 0 to 200 nM. The
effect of added H2O2 (0 – 500 nM), pH (6.0 – 8.5) and NaHCO3 (2 – 9 mM) on the Fe(II) rate constants was
studied at Cu(II) levels (0 – 200 nM). To understand the competition between Fe and Cu, the reduction of Cu(II)
to Cu(I) was also studied as a function of oxygen (air-saturated and anoxic seawater), Fe(II) (0 – 200 nM) and
H2O2 (0 – 300 nM). The Fe(II) oxidation was accelerated by the presence of Cu(II) and Cu(I). This acceleration
has been explained by the redox coupling between Fe and Cu, competition for different inorganic species
(hydroxyl and carbonate groups studied independently) and by the formation of Fe-Cu particles (cupric or cuprous
ferrite). Superoxide played a key role in the oxidation rate of Fe(II) in the presence of Cu(II). The presence of
Fe(II) caused a greater reduction of Cu(II) to Cu(I). This is directly related to the levels of oxygen, Fe(II) and
H2O2 concentrations. The presence of Fe(II) produced a rapid formation of Cu(I) in the first 2 – 3 min of reaction.
These experimental results demonstrated that the presence of Fe and Cu strongly affected the inorganic redox
chemistry of both metals in UV-treated seawater.
INTRODUCTION
The effect of copper on the Fe chemistry has been studied
[1, 2, 3, 4, 5]. These authors found that the redox reaction
of both metals can be affected by the redox reactions of the
other. These results showed a true catalytic effect of copper
on the Fe(II) oxidation in acidic solutions [1]. This was
caused by the radical hydrosuperoxide, HO2•, which reacts
with Fe(II) according to the Haber-Weiss mechanism. The
same authors also reported that the effect of Cu(II) on
Fe(II) kinetics is a function of the initial concentrations of
both Fe and Cu. Sayin [3] suggested that copper ions take
part in an oxidation-reduction cycle of Fe(II) and act as a
catalyst in the oxidation of the iron in biotite. They
demonstrated that the oxidation of iron increased in the
presence of Cu. Sedlak and Hoigne [4] studied the role of
copper in the redox cycling of iron in atmospheric waters
and also demonstrated that the catalytic effect of copper
altered the production of radicals during the oxidation.
Matocha et al. [5] showed that, under anoxic and acidic
conditions, the reduction of Cu(II) to Cu(I) by dissolved
Fe(II) was rapid and generally completed in the first 1-2
min and was also affected by the presence of chloride ions.
The oxidation of Fe(II) by Cu(II) is thermodynamically
possible by considering the speciation of these metals in
solution [5]. In this manuscript, we focused on the
inorganic interactions between Fe and Cu in seawater, in
terms of redox reactions. The oxidation rate of Fe(II) in
seawater at nanomolar levels is studied in the presence of
both Cu(I) and Cu(II) to levels of 200 nM. The effect of pH
(6.0 – 8.5), bicarbonate concentrations (2 – 9 mM) and
H2O2 (0 – 500 nM), superoxide (after superoxide dismutase
treatment) in the presence of copper was also studied. The
effect of chloride on the Cu(I) chemistry is minimized here
due to the constant concentration of chloride in seawater.
The effect of Fe on the Cu redox chemistry has been
studied under different experimental conditions, such as a
function of Fe(II) concentration (0 – 200 nM), under airsaturated and anoxic conditions, together with the effect of
H2O2 levels (0 – 300 nM). These results should improve
our knowledge of the importance of these metal
interactions in natural waters and improve the kinetic
models for the oxidation of Fe(II) and Cu(I) in seawater.
MATERIALS AND METHODS
The Fe(II) and Cu(I) concentrations were determined
spectrophotometrically using the ferrozine method [6, 7]
and a modified version of the bathocuproine method [8, 9]
respectively. The ferrozine and Fe(II) form a peak at 562
nm. Bathocuproine disulfonate salt and Cu(I) form a peak
at 484 nm [8, 9]. Kundra et al. [10] demonstrated the
formation of Cu(I)-ferrozine at 470 nm. All our
experimental data were corrected by the non-absorbing
baseline at 700 nm.
Fe(II) and Cu(I) were measured at nanomolar levels by
using a 5 m long waveguide capillary flow cell (World
Precision InstrumentsTM) connected to the UV-Vis
detector USB2000 (Ocean OpticsTM). The light used was
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a halogen light source (HL-2000-FHSA from Mikropack).
The capillary flow cell and the UV detector were connected
using 400 µm optical fiber. The spectra were recorded
using the OOIBase32 software by Ocean Optics. The
sample was introduced in the column using a peristaltic
pump (EXPETEC Perimax 12) with a flux of 1 mL min-1.
RESULTS AND DISCUSSION
This study demonstrates how the redox chemistry of Fe is
affected by Cu in seawater. Cu(II) and Cu(I) accelerate the
Fe(II) oxidation rate in seawater. The key role of O2•- on
the Fe(II) rate in the presence of Cu(II) has been
demonstrated. The increase in the oxidation rate of Fe(II)
can be due to the added Cu, and to the competition of Fe
and Cu interaction with hydroxyl, H2O2, O2•- and carbonate
groups. These ROS reactions are not enough to explain the
Fe-Cu redox interaction and the Fe-Cu redox pair reaction,
and the formation of Fe-Cu particles (cupric and cuprous
ferrite) should be considered.
The effect of Fe(II) on the reduction of Cu(II) was also
studied. Cu(II) was rapidly reduced to Cu(I) in the presence
of Fe(II) in seawater under air-saturated and anoxic
conditions. The concentration of Cu(I) formed was affected
by the concentrations of both Fe(II) and H2O2 in seawater.
This study demonstrates that the oxidation of Fe(II) in the
presence of inorganic Cu increased the oxidation rate in
seawater. For Cu(II), the presence of Fe(II) invoked a
strong and rapid reduction to Cu(I) that underwent
oxidation, under air-saturated conditions. Sayin [3]
suggested an overall reaction to describe the Fe-Cu
interaction in acidic solutions (Equation 1 – 4). These
reactions can be also applied to the experimental conditions
considered in the current investigation according to the
experimental results.
4Fe(II) → 4Fe(III) + 4e−
4Cu(II) + 4e− → 4Cu(I)
4Cu(I) → 4Cu(II) + 4e−
4H+ + O2 + 4e− → 2H2 O
(1)
(2)
(3)
(4)
4Fe(II) + 4H+ + O2 ↔ 4Fe(III) + 2H2 O
(5)
with the overall reaction (Equation 5):
where Fe(II), Fe(III), Cu(I), and Cu(II) represent the free or
inorganic species in natural waters. The formation of Cu-Fe
particles (cupric or cuprous ferrite) cannot be discriminated
as well as the Fe(III) and Cu(I) interaction which was
described as a quick reaction by Sedlak and Hoigne [4]. In
order to summarize the effect of Cu on the Fe(II) oxidation
in seawater we proposed the layout of the interaction (Fig.
1) that is a modified version from Sedlak and Hoigne, [4],
where Fe(II) is rapidly oxidized to Fe(III).
Fig. 1. Layout of the interaction between Fe and Cu during
the redox reactions in seawater (modified from Sedlak and
Hoigne [4]).
ACKNOWLEDGMENTS
This study received financial support from the Project
CTM2010-19517-MAR and CTM2014-52342-P given by
the Ministerio de Economía y Competitividad from Spain.
REFERENCES
1 - Stumm W & Lee GF, 1961. Kinetic product of ferrous
iron. Ind. Eng. Chem., 53: 143-146.
2 - Parker OJ &Espenson JH, 1969. Reactions involving
copper(I) in perchlorate solution. A kinetic study of the
reduction of iron(III) by copper(I). Inorg. Chem., 8(7):
1523-1526.
3 - Sayin M, 1982. Catalytic action of copper on the
oxidation of structural iron in vermiculitized biotite. Clay
Miner., 30(4): 287-290.
4 - Sedlak DL & Hoigne J, 1993. The role of copper and
oxalate in the redox cycling of iron in atmospheric waters.
Atmos. Environ. - Part A General Topics, 27(14): 21732185.
5 - Matocha CJ, Karathanasis AD, Rakshit S & Wagner
KM, 2005. Reduction of copper(II) by iron(II). J. Environ.
Qual., 34: 1539-1546.
6 - Violler E, Inglett PW, Hunter K, Roychuodhury AN &
Cappellen P, 2000. The ferrozine method revisited:
Fe(II)/Fe(III) determination in natural waters. App.
Geochem. 15, 785-790.
7 - Santana-Casiano JM, Gonzalez-Davila M & Millero FJ,
2005. Oxidation of nanomolar levels of Fe(II) with oxygen
in natural waters. Environ. Sci. Tech. 39(7), 2073-2079.
8 - Moffett J, Zika R & Petasne R, 1985. Evaluation of
bathocuproine for the spectrophotometric determination of
copper (I) in copper redox studies with applications in
studies of natural-water. Anal. Chim. Acta. 175, 171−179.
9 - Gonzalez-Davila M, Santana-Casiano JM, Gonzalez
AG, Perez N & Millero FJ, 2009. Oxidation of copper (I)
in seawater at nanomolar levels. Mar. Chem. 115(1), 118124.
10 - Kundra SK, Katyal M & Singh RP, 1974.
Spectrophotometric determination of copper(I) and
cobalt(II) with ferrozine. Analytical Chemistry, 46(11):
1605-1606.
80
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The science of ocean predictions and operational oceanography: the new
science paradigm
Nadia Pinardi1, Francesco Trotta1, Giovanni Coppini2, Emanuela Clementi3, Claudia
Fratianni3 and Ivan Federico2
1
Department of Physics and Astronomy, Alma Mater Studiorum University of Bologna, Bologna (IT)
Centro EuroMediterraneo sui Cambiamenti Climatici, Ocean Predictions and Applications, Lecce (IT)
3
Istituto Nazionale di Geofisica e Vulcanologia, Bologna section, Bologna (IT)
2
ABSTRACT
The science of ocean predictions has started in the eighties and it has rapidly advanced thereafter due to satellite
sea level data availability and increasingly accurate numerical ocean models. The last ten years have seen great
advances in this new sector of oceanography: numerical ocean models that resolve the mesoscales were
implemented from the global ocean to the regional seas (scales of few km), data assimilation schemes capable to
assimilate frequent ocean profiles and satellite data were developed, coupling between eulerian hydrodynamics
and surface waves was started to better resolve the surface currents and nesting of unstructured ocean models
allows to forecast properly the coastal, tidal and baroclinic currents at the resolution of few hundred meters. The
predictability limit for ocean short term forecasts is of the order of several days, depending on the ocean variable
and the accuracy of the atmospheric forcing forecast.
The possibility to produce analyses and forecasts at the ocean mesoscales is underpinning our new capacity to
understand ocean dynamics and investigate the climate variability of the ocean: ocean forecasting models in fact
are also used to produce reanalyses that allow an accurate reconstruction of the mean and eddy components of
the flow field and their relationship, describe wáter mass formation processes and study the ocean dynamics at
unprecedented resolution and accuracy. On the other hand, the access to open data from the operational services
and in particular the Copernicus Marine Enrvironment Monitoring Service (http://marine.copernicus.eu/), allows
to develop new ocean applications for the blue economy that were unthinkable few years ago, among others,
accurate search and rescue decision support systems, oil spill forecasting and hazard mapping, efficient ship
routing and good environmental status marine indicators.
81
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Distribution and ecological risk assessment of legacy and emerging organic
pollutants in coastal marine sediments from the Atlantic coast (Andalusia,
SW Spain)
Marina G. Pintado-Herrera1, Tatiane Combi2, Carmen Corada-Fernández1, Eduardo GonzálezMazo1 & Pablo A. Lara-Martín1
1
Physical Chemistry Department, Faculty of Marine and Environmental Sciences, University of Cadiz. Campus de Excelencia
Internacional del Mar (CEI-MAR). Cadiz, Spain, 11510
2
Interdepartmental Centre for Environmental Science Research, University of Bologna, Via San Alberto 163, 48123 Ravenna,
Italy
ABSTRACT
Contamination of aquatic systems by no longer used but very persistent compounds (e.g., organochlorine
pesticides) and newly detected chemicals, such as personal care products (PCPs), represents a raising concern. We
have aimed in this work to carry out one of the first comparisons of both types of contaminants, legacy and
emerging, in two coastal systems (Cadiz Bay and Huelva Estuary). A wide range of analytes were selected to this
end, including hydrocarbons, UV filters, fragrances, antimicrobials… Analysis of surface sediments revealed the
occurrence of 46 out of 97 chemicals, and that most of them were predominantly accumulated in depositional
areas with high organic carbon content. Polycyclic aromatic hydrocarbons (PAHs), fragrances (e.g., OTNE), UV
filters (e.g., octocrylene), and nonylphenol had the highest concentrations (up to 1098, 133.5, 72 and 575 ng g-1,
respectively). Several inputs were detected, from atmospheric deposition after combustion to wastewater
discharges and recreational activities. However, an environmental risk assessment performed for those chemicals
for which ecotoxicological data was available, indicates that legacy contaminants still pose the highest potential
risk towards benthonic organisms (individual hazard quotients up to 580).
INTRODUCTION
Up until the end of the 1990s most of the environmental
monitoring on xenobiotic organic compounds was devoted
to persistent organic pollutants (POPs) and other priority
contaminants due to their well-known persistence,
bioaccumulation potential and toxicity. Although the
production and use of many of them (e.g., polychlorinated
biphenyls, or PCBs) are nowadays banned or restricted in
most countries (“legacy contaminants”), knowing the
actual distribution of these chemicals is relevant since they
are still detected in the environment and their advisable
levels are set by several international laws. More recently,
there has been a rising interest for identifying and
screening new organic synthetic compounds in the
environment, the so-called “emerging contaminants”,
which has been possible due to the development of new
analytical techniques. One of the main groups of emerging
contaminants is personal care products (PCPs), substances
that are widely consumed by the society (e.g., surfactants,
fragrances, UV filters, antimicrobials, etc.) and
continuously introduced in the environment mainly through
the effluents of wastewater treatment plants (WWTPs). The
knowledge on the distribution and fate of organic
contaminants, both legacy and emerging compounds, in the
particulate phase (e.g., sediments, suspended solids, and
soils) is more limited than in the aqueous phase as their
bioavailability and, hence, toxicity, is severely reduced. In
that sense, this work is focused on investigating and
comparing the spatial distribution of a wide range of both
legacy and emerging contaminants in coastal systems from
SW Spain (Cadiz Bay and Huelva Estuary). As well as
performing an initial assessment of their potential risk for
benthic organisms. For this purpose, 97 structurally diverse
compounds have been monitored, including PAHs, PCBs,
several types of pesticides (e.g., organochlorines,
organophosphorus, and pyrethroids), organophosphate
flame retardants, antimicrobials, nonylphenol, fragrances,
and UV filters.
STUDY AREA AND EXTRACION PROCEDURE
Two different coastal systems from the Atlantic coast of
Andalusia (SW Spain) have been selected: the Cadiz Bay
and the Huelva Estuary (Figure 1). Both areas comprise
salt-marsh environments surrounded by coasts and several
towns, accounting for a total of about 600 000 and 260 000
inhabitants, respectively.
Surface sediment samples (0-5 cm depth) were taken at 48
samplings stations in Cadiz Bay area and 18 in Huelva
Estuary using a Van Veen grab. Duplicates of each sample
were extracted and analyzed using the methodology
82
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previously developed by Pintado-Herrera et al. [1].
Extraction of analytes from sediment samples was achieved
by pressurized liquid extraction (PLE). Separation,
identification and quantification of target compounds were
performed on a gas chromatography-tandem mass
spectrometry.
DATA ANALYSIS AND RISK ASSESSMENT
An interpolation method was applied to estimate the
concentration of target compounds in surface sediments
from both sampling areas, by using the ArcGIS 10
software. Then, mass inventories for different classes of
contaminants were calculated taking into account the
contaminant concentration in sediment, the thickness of
sediment sampled and the average density of the dry
sediment particles. Additionally, a risk evaluation was also
performed for those target compounds that were detected in
sediments and for which ecotoxicological data on
benthonic species was available. Measured environmental
concentrations (MEC) were derived from the data
presented in this manuscript whereas the predicted non
effect concentrations (PNEC) were taken from literature.
Using both MEC and PNEC values, different hazard
quotients (HQ) were calculated.
RESULTS AND DISCUSSION
Forty six out of 97 analytes were found in samples from
Cadiz Bay, whereas 43 were detected in Huelva Estuary.
Regarding the Cadiz Bay, the main sources of organic
contaminants that affect this area are wastewater discharges
from local WWTPs and aquatic recreational activities
during the touristic season. There are also some active
fishing ports, a military naval station and some heavy
industries (shipyards and a plane factory). On the other
hand, Huelva Estuary is characterized by receiving acidic
fluvial water discharges (pH <4) with very high
concentrations of heavy metals through the Tinto and Odiel
river system. Other than mining, there is a very important
industrial area at the junction of the two rivers. Occurrence
of sewage-derived contaminants from local WWTPs and
pesticides from strawberry fields is also expected.
Roughly, most of the contaminants followed similar
distribution patterns (Figure 1), presenting the highest
concentrations in the western part of the inner bay (Cadiz)
and along Tinto River (Huelva). Nevertheless, different
contamination sources were identified depending on the
chemical and the sampling site (e.g., ports for PAHs,
recreational activities for UV filters, and wastewater
discharges for fragrances). Regarding the sources of PAHs
in both aquatic environments, all the results showed similar
trends and pointed out petroleum and biomass combustion
(e.g., coal) as the common sources for PAHs in these areas.
Some compounds, such as the UV filters octocrylene or
EHMC, have been identified as ubiquitous in both areas.
Their relative higher levels in sediments compared to other
UV filters can be explained on the basis of their higher
hydrophobicity and their extensive use in sunscreen
formulations.
To evaluate the extent of contamination, mass inventories
were calculated in Cadiz Bay and Huelva Estuary (Figure
2), showing that PAHs was the most abundant class of
contaminants in terms of mass (528 kg and 411 kg,
respectively), followed by nonylphenols (408 kg in Cadiz
Bay). In general, we could also observe that sediments with
high organic carbon content and fine-grain (<63µm) had
higher levels of many of the target analytes (PAHs, PCBs,
fragrances …).
Fig. 1. PAH spatial distribution: 1) Cadiz Bay and 2)
Huelva Estuary.
Finally, results from a preliminary environmental risk
assessment indicated that there is a potential risk for some
compounds (mainly DDE, PCBs and PAHs) to cause
biological adverse effects. There is, however, lack of
reliable information to perform an accurate prediction of
the risk to human health and aquatic organisms after
decades of continuous exposure to random combinations of
low levels of both regulated and emerging products.
Fig. 2. Mass inventories of the analyzed groups in both
study areas.
ACKNOWLEDGEMENTS
This study has been carried out with the support of two
Spanish regional research projects (RNM 5417 and RNM
6613), and with the help of a grant from the Spanish
Ministry of Education, Culture and Sport.
REFERENCES
1 - M.G. Pintado-Herrera, González-Mazo, E., LaraMartín, P.A, 2016. In-cell clean-up pressurized liquid
extraction and gas chromatography-tandem mass
spectrometry determination of hydrophobic persistent and
emerging organic pollutants in coastal sediments. J.
Chromatogr. A 1429:107-118.
83
(2016)
Abnormal mortality of octopus after a storm water event: accumulated lead
and stable lead isotopes as fingerprints
Joana Raimundo1,2, Francisco Ruano1, João Pereira1, Mário Mil-Homens1,2, Pedro Brito1,2,
Carlos Vale2 & Miguel Caetano1,2
1
IPMA - Portuguese Institute of Sea and Atmosphere, Rua Alfredo Magalhães Ramalho, 6, 1495-006 Lisbon, Portugal
CIIMAR, Marine and Environmental Research Center, Rua dos Bragas, 289, 4050-123 Porto, Portugal
2
ABSTRACT
Octopus vulgaris is a sedentary organism that inhabits coastal waters being subjected to anthropogenic
compounds from terrestrial origin. Digestive gland accumulates high levels of Pb and other contaminants as a
result of its physiological function in the digestive process. Lead concentration in coastal environments results
from weathering, industrial and domestic discharges, and atmospheric deposition. Since stable Pb isotopic
composition is not affected by kinetic processes occurring between source and sink, its signature in marine
sediments has been used to identify the influence of the different sources. However, lead isotopic signature in
organisms has been poorly explored in environmental sciences. After a short period of a heavy rainfall, hundreds
of octopus died in Portuguese coastal areas. Levels of Pb and ratios of its stable isotopes were determined in the
digestive gland of stranded octopus and compared to individuals caught alive and to sediments and urban runoff
material. Octopus digestive gland reflected Pb concentrations in sediments, with enhanced levels in specimens
from Cascais. Pronounced augment (up to one order of magnitude) of this element was registered in the digestive
gland of stranded octopus. Stranded octopus presented lower Pb isotope signatures than the alive ones but closer
to the signatures obtained in the runoff material. The obtained results clearly evidence that octopus mass stranding
was related to strong rainfall and runoff. Pb isotopic signature was an adequate proxy to determine the causeeffect relation.
INTRODUCTION
Urban water runoff is one of the major sources of
pollutants to receiving waters [1-3]. In urban environments,
anthropogenic Pb originates from a variety of sources
including vehicular dust, leaded paint and industrial
emissions. The study of stable Pb isotopes provides a
powerful tool in tracing Pb sources. The ratios between
isotopes provide an identification of different Pb sources
which have distinct isotopic signatures [4, 5]. The Pb
isotope ratios vary with local geology and proximity to
anthropogenic inputs, as well as temporally as pollution
sources change.
Cephalopods represent an essential link in marine trophic
chains. O. vulgaris have a short life span, high metabolic
rates and inhabit at coastal waters. Several studies have
highlighted that octopus digestive gland can reflect
environmental contamination [6-8]. In some cases,
geographical variations of metal availability can overcome
the biological differences [7]. Concentrations of Pb have
contrasting geographic patterns in digestive gland of
specimens collected in the Portuguese coast [8-10].
After a period of a heavy rainfall, hundreds of octopus died
in two Portuguese coastal areas adjacent to rivers with high
flow. The aim of this study was to test Pb isotopic
signature as a tool to trace the rapid and high input of
freshwater/runoff material as the primary cause of octopus
mortality. This hypothesis was tested by: i) histological
alterations in octopus tissues; and ii) determination of Pb
concentrations and Pb stable isotopes in digestive gland,
sediments and urban runoff material.
MATERIAL AND METHODS
Seventy five common octopuses, Octopus vulgaris, were
collected in two areas, Matosinhos (N of Portugal) and
Cascais (nearby Lisbon), of the Portuguese coast. Octopus
were captured in two contrasting conditions: (i) living
specimens caught in November 2009 (hereafter “alive”);
(ii) dead specimens collected in January 2010, after a
period of heavy rain and runoff (hereafter “stranded”).
Digestive gland was totally removed, freeze-dried,
grounded and homogenised. Pieces of mantle, arms and
digestive gland were collected from stranded and alive
octopuses and used for histological analysis, following the
protocol adopted by the IPMA laboratory of pathology for
mollusks.
The top 5-cm layer of sediments were sampled in 2009 at
Matosinhos (n=5) and Cascais (n=4) with a Van-Veen
grab. The urban runoff-derived material resulted from
heavy rain periods in January 2010 was collected from two
roadway gully pots. Sediments and urban runoff material
were oven-dried, sieved (2-mm mesh) and grounded.
84
(2016)
1.2000
Runoff
Alive Casc
1.1400
1.1200
0.4700
0.4750
206 Pb/208 Pb
206
High precipitation event have occurred just before the
stranded octopus were found. The average rainfall changed
from 43 mm to 216 mm within two weeks. Under these
conditions, the death of the octopus was presumably related
to the rapid decrease of salinity, since cephalopods show a
high sensitivity to changes in salinity conditions.
The main tissue structures of stranded octopus were
severely compromised, showing lesions in the collagen
fibers as well as in the cells of the digestive gland. This
suggests that death was likely due to multiple organ failure,
related to hypertrophy and presence of exudates that
correlate with the extreme environmental alterations. The
accumulation of fluid in octopus tissues leading to their
death may concomitantly have conduced to the input of
other solutes.
The geographical differences observed for Pb
concentrations in the digestive gland of octopuses, higher
in Cascais, are in line with the distinct levels registered in
the sediments from the two areas. High Pb concentrations
were observed in stranded specimens, which were up to
one order of magnitude above the levels reported for alive
octopus captured in the Portuguese coast [14-16,
unpublished data]. The sharp increase of Pb
bioaccumulation was presumably associated with the
runoff period. Alive octopus showed high Pb signatures,
while stranded organism showed lower isotopic ratios
closer to the urban runoff material (Figs. 1 and 2).
1.2200
Runoff
Sed Mat
Alive Mat
Strand Mat
206 Pb/207 Pb
1.2000
1.1800
1.1600
1.1400
1.1200
1.1000
0.4650
0.4700
0.4750
0.4800
206 Pb/208 Pb
0.4850
0.4900
Fig. 1. Relationships between 206Pb/207Pb and 206Pb/208Pb
ratios in urban runoff material, sediments, alive and
stranded O. vulgaris captured in Matosinhos (Mat).
Strand Casc
1.1600
1.1000
0.4650
RESULTS AND DISCUSSION
Sed Casc
1.1800
206 Pb/207 Pb
Samples of digestive gland were digested according to
[11]. For sediments and runoff material, two mineralization
procedures were used according to the methods by [12, 13].
Total Pb concentration and stable Pb isotopes (206Pb, 207Pb
and 208Pb) were determined using a quadrupole ICPMS.
The precision and accuracy of Pb concentration and stable
isotope composition was determined through analysis of
certified reference materials and using 115In as internal
standard.
0.4850
0.4800
207
206
Fig. 2. Relationships between Pb/ Pb and Pb/208Pb
ratios in urban runoff material, sediments, alive and
stranded O. vulgaris captured in Cascais (Casc).
The obtained results suggest that octopus mass strandings
were related to strong rainfall and runoff. Moreover, Pb
isotopes results showed to be a good proxy to determine
runoff inputs in the environment and in biological matrices
and in this case provided useful forensic evidence in the
identification of the cause-effect.
ACKNOWLEGDMENTS
Joana Raimundo acknowledges the pos-doctoral grant by
FCT (SFRH/BPD/91498/2012), and João Pereira the
support of COST Action FA1301.
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1 - Sansalone JJ & Buchberger SG, 1997. J. Environ. Eng
ASCE,123: 134-143.
2 - Barrett ME, Irish Jr LB, Malinia Jr JF & Charbenuea
RJ, 1998. J. Environ. Eng ASCE,124: 131-137.
3 - Davis AP, Shokouhian M & Ni S, 2001. Chemosphere,
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4 - Labonne M, Othman DB & Luck J-M, 1998. App.
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5 - Labonne M, Othman DB & Luck J-M, 2001. Chem.
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Environ., 340: 113-122.
11 - Ferreira A, Cortesão C, Castro O & Vale C, 1990. Sci.
Total Environ., 97/98: 627-639.
12 - Loring D & Rantala R, 1990. Tech. Mar. Environ. Sci.,
9: 1-13
13 - Caetano M, Fonseca N, Cesário R & Vale C, 2007.
Sci. Total Environ., 380:84-92.
14 - Raimundo J, Vale C, Duarte R & Moura I, 2008. Sci.
Total Environ., 390: 410-416.
15 - Raimundo J, Vale C, Caetano M, Cesário R & Moura
I, 2009. Aquat. Biol., 6: 25-30.
16 - Raimundo J, Costa PM, Vale C, Costa M & Moura I,
2010. Comp. Biochem. Physiol. C, 152: 139-146.
85
(2016)
Biomarkers response in Scrobicularia plana following exposure to
polystyrene microplastics
Francisca G. Ribeiro, Nélia C. Mestre, Cátia Cardoso, Tainá Fonseca, Thiago L. Rocha,
Maria Fonseca, Beatriz Pereira, Maria João Bebianno
CIMA, Centre for Marine and Environmental Research, University of Algarve, Campus de Gambelas, 8005-135 Faro,
Portugal
ABSTRACT
Nowadays there is an increasing resilience of plastics as an everyday item. With the rapid increase in their
production and spread, the plastic debris are accumulating in the marine environment where they are fragmented
into smaller pieces. One of the most produced polymer, and accordingly, more common in the marine
environment is the polystyrene (PS). Ranges of organisms, especially invertebrates, are vulnerable to the
exposure of microparticles. However, the impacts of microplastics (< 5mm) in the marine systems are poorly
understood. The aim of this study is to assess the toxicity of PS microplastics in different tissues of the peppery
furrow shell Scrobicularia plana. Clams were exposed to 1 mg L-1 of PS microplastics (20 µm) for 14 days,
followed by a 7 days depuration. Microplastics accumulation in gills and digestive gland, and their effects on a
battery of biomarkers response: oxidative stress (catalase, superoxide dismutase, glutathione-S-transferase and
glutathione peroxidase), oxidative damage (lipid peroxidation), neurotoxicity (acetylcholinesterase activity) and
genotoxicity (comet assay to evaluate DNA damage) were assessed to select the most appropriate biomarker to
evaluate microplastics effects.
INTRODUCTION
In the contemporary society, plastics have acquired a
fundamental importance to commercial, industrial and
medical applications. Polypropylene (PP), polyethylene
(PE) and polystyrene (PS) are the most produced polymers,
and thus, more common in the marine environment [1].
Recently, ubiquitous microscopic particles were identified
in the marine environment - the microplastics. Currently,
they are defined as particles that are less than 5 mm in
diameter, according to the National Oceanic and
Atmospheric Administration of the United States of
America [2]. They are an emerging marine contaminant
and, to date, have been found in many habitats and in the
system of a variety of marine and freshwater species. Thus,
it is important to understand its distribution in the marine
environment and the implications on habitats, biodiversity
and health of the marine species [1]. The biological effects
in organisms depend on the size of the microplastics
whereas, the smaller the size, the greater the effect to be
accumulated at the cellular level. Despite concerns related
to the ingestion, the effects of microplastics in populations
and its implications for the food chain are not well known.
Many adverse effects of microplastics in a lot of marine
species have been described in the literature, such as:
reduction of the feed rate, physical damage due to
accumulation, induction of oxidative stress, effects on
reproduction,
decreased
neuroprotective
activity,
development of pathologies and even mortality [3, 4].
Marine invertebrates are particularly susceptible to
microplastics, because of their size and feeding strategy.
Scrobicularia plana was used as the model to study
microplastics toxicity. Since the microplastics mode of
action and biological risk are not yet very clear, the
potential ecotoxicological risk need to be assessed.
Reactive oxygen species (ROS) can damage lipids,
proteins, and DNA. Thus, stress oxidative enzymes play an
important role as a defense response and form an important
component of the antioxidant response [5]. In the present
work the impact of Polyestirene (Ps) was assessed, in gills
and digestive gland of S. plana by the response of oxidative
enzimes (catalase, superoxide dismutase, glutathione Stransferase and glutathione peroxidase)
MATERIAL AND METHODS
Scrobicularia plana bivalve species (38 ± 5 mm of shell
length) were collected in Cabanas de Tavira, Algarve,
Portugal (N 37º7'59.75 '' W 7 36'34.95 ''). Live animals
were transferred to the laboratory and inserted in 20 liters
glass aquaria (n=6), with constant aeration. They were
acclimated during 7 days and exposed to PS microplastics
for 14 days, followed by 7 days of depuration. The aquaria
were divided into two groups: control and exposed
(concentration of PS of 1 mg L-1), in triplicate. The
monodisperse PS microplastics (diameter = 20 µM) were
obtained from Sigma-Aldrich (Germany). A stock solution
(100 mg L-1) was prepared in ultrapure water (18 MΩ/ cm)
and, before every renewal, sonicated for 30 minutes
86
(2016)
(Ultrasonic bath VWR International, 230 V, 200 W, 45
kHz frequency). The water was changed every 24 hours,
with subsequent addition of 1 ml of the microplastics stock
solution. Clams were not fed and no significant differences
in mortality, between treatments, were detected during the
exposure. Clams were collected after 0, 3, 7 and 14 days of
exposure, and after the 7 days of depuration. Gills,
digestive gland, and remaining tissues (mantle, foot, and
adductor muscles) were dissected and stored at -80 ºC, until
further use.
To assess the physiological status of control and exposed
clams to microplastics during the experiment, soft tissues
and shells were weighted and the condition index (CI) was
determined as the percentage (%) of the ratio between
drained weight of the soft tissues (g) and total weight (g).
Antioxidant enzymatic activities were measured in the gills
and digestive glands cytosolic fractions of clams, from
control groups and exposed to microplastics. Total protein
concentration was determined in the cytosolic fraction
following the Bradford method [6], using bovine serum
albumin as a standard. SOD activity was determined in the
cytosolic fraction by the reduction of cytochrome c by the
system xanthine oxidase/hypoxanthine at 550 nm [7]. CAT
activity was determined by measuring the consumption of
hydrogen peroxide (H2O2) at 240 nm [8]. GPx activity was
measured through NADPH oxidation in the presence of
excess glutathione reductase, reduced glutathione and
hydroperoxide as substrate, at 340 nm [9]. GST activity
was determined by measuring the conjugation of l-chloro2,4-dinitrobenzene (CDNB) with reduced glutathione
(GSH), by an increase in absorbance at 340 nm [10].
RESULTS AND DISCUSSION
No significant changes were obtained in the condition
index, between unexposed and exposed organisms, during
the accumulation (control: 33.05 ± 4.76 %; microplastics:
31.53 ± 5.30 %; p > 0.05) and the depuration periods
(control: 31.31 ± 4.58 % microplastics: 31.83 ± 4.72 %; p
>0.05), indicating that the organisms were in good health
during the experiment.
Regarding the gills, results showed that the exposure to
microplastics significantly induce an increase in SOD
activity, (p<0.05) after 7 and 14 days. However, GPx
activity only increase in clams exposed to microplastics on
the 3rd day of exposure, and GST activity only on 14 day of
exposure. During the depuration there was only a decrease
in SOD activity (p<0.05).
In the digestive gland, results showed an increase on SOD
activity on the 14th of accumulation period. Moreover, GPx
activity was significantly increase on day 3 followed by a
decrease on the 14th (p<0.05), while GST activity
significantly decreased in clams exposed to microplastics
after 3 days of exposure (p<0.05). Followed the depuration
period, SOD increased its activity, while GPx and GST
activity were inhibited (p<0.05). Previous studies reported
that microplastics can cause oxidative stress responses.
These results also indicated like previous data an increase
in oxygen and nitrogen based reactive species, that are the
first inflammatory response, induced by polystyrene
nanoparticles, in the hemocytes of Mytillus [11] and an
inhibition for Se-dependent glutathione peroxidases and
catalase in the digestive gland of mussels exposed to
polystyrene and polyethylene [12].
Both gills and digestive gland seem to be susceptible to
oxidative stress, since enzimatic activities were noticiable
in both tissues. So, we can predict that probably PS
microplastics induce ROS and can cause toxic effects in
clams. Nevertheless, microplastics toxicity needs to be
more fully addressed, by combining the results with other
biomarkers.
REFERENCES
1 - Lusher, A., 2015. Microplastics in the marine
environment: distribution, interactions and effects. Marine
anthropogenic litter. Springer, pp. 245-307.
2 - NOAA, 2015. National Oceanographic Administration
Service
3 - Browne, M. A., Dissanayake, A., Galloway, T. S.,
Lowe, D. M., Thompson, R. C., 2008. Ingested
microscopic plastic translocates to the circulatory system of
the mussel, Mytilus edulis (L.). Environmental science &
technology 42, 5026-5031.
4 - Wright, S. L., Thompson, R. C., Galloway, T. S., 2013.
The physical impacts of microplastics on marine
organisms: a review. Environmental Pollution 178, 483492.
5 - Cid, A., Picado, A., Correia, J. B., Chaves, R., Silva, H.,
Caldeira, J., Diniz, M. S. (2015). Oxidative stress and
histological changes following exposure to diamond
nanoparticles in the freshwater Asian clam Corbicula
fluminea (Müller, 1774). Journal of Hazardous Materials,
284, 27-34.
6 - Bradford, M. M., 1976. A rapid and sensitive method
for the quantitation of microgram quantities of protein
utilizing the principle of protein-dye binding. Analytical
biochemistry 72, 248-254.
7 - McCord, J. M., Fridovich, I., 1969. Superoxide
dismutase an enzymic function for erythrocuprein
(hemocuprein). Journal of Biological chemistry 244, 60496055.
8 - Greenwald, R. A., 1987. Handbook of methods for
oxygen radical research. Free Radical Biology and
Medicine, 161.
9 - Lawrence, R. A., Burk, R. F., 1978. Species, tissue and
subcellular distribution of non Se-dependent glutathione
peroxidase activity. The Journal of nutrition 108, 211-215.
10 - Habig, W. H., Pabst, M. J., Jakoby, W. B., 1974.
Glutathione S-transferases the first enzymatic step in
mercapturic acid formation. Journal of biological
Chemistry 249, 7130-7139.
11 - Canesi, L., Ciacci, C., Bergami, E., Monopoli, M.,
Dawson, K., Papa, S., Canonico, B., Corsi, I., 2015.
Evidence for immunomodulation and apoptotic processes
induced by cationic polystyrene nanoparticles in the
87
(2016)
hemocytes of the marine bivalve Mytilus. Marine
environmental research, 111, 34-40.
12 - Avio, C. G., Gorbi, S., Milan, M., Benedetti, M.,
Fattorini, D., d'Errico, G., Pauletto, M., Bargelloni, L.,
Regoli, F., 2015. Pollutants bioavailability and
toxicological risk from microplastics to marine mussels.
Environmental
Pollution
198,
211-222.
88
(2016)
Influence of seawater carbonate system in phenotypic plasticity response to
ocean acidification of the mollusc Scurria araucana along the Chilean coast
Araceli Rodríguez-Romero1,3, Tania Opitz1, Samanta Benítez2,3, Laura Ramajo1,2, Nelson A.
Lagos2,3, Marcos A. Lardies1,3
1
Facultad de Artes Liberales & Facultad de Ingeniería y Ciencias, Universidad Adolfo Ibáñez, Peñalolén, Santiago, Chile
Facultad de Ciencias, Universidad Santo Tomas, Ejército 146, Santiago, Chile
3
Center for the Study of Multiple-drivers on Marine Socio-Ecological Systems (MUSELS).
2
ABSTRACT
The rapid increase of surface ocean concentrations of CO2, due to rising anthropogenic CO2 emissions is expected
to exert negative impacts on a wide range of marine fauna. One of the major challenges is to understand how
marine organisms will respond to ongoing rapid ocean acidification (OA). The main aim of this study is to predict
how organisms will respond differentially to OA due to the natural variability of pCO2 in their respective habitats.
For that, we investigate the geographic variation in seawater carbonate system and their effects in the
physiological phenotype traits using both, a field and a laboratory approaches in three populations of the mollusc
Scurria arauacana, Huasco, Talcaruca and Los Molles localized at both sides of one of the most important
biogeographic break (i.e. originated by permanent upwelling) on the Chilean coast. Individuals of S.araucana
populations were collected during two years in the selected localities and transported to the laboratory. Once at
laboratory, organisms were characterized in terms of shell length, total buoyancy and dry tissue weight, shell
composition, metabolic and heart . After acclimation period individuals of populations from Talcaruca and Los
Molles were exposed to three CO2 treatments (400 ppm, 750 ppm and 1300 ppm) for 30 days. Metabolic rates and
buoyancy weight were determined at 10, 20 and 30 days. Our results shows that carbonate content of shells and
tissue of organisms differ significantly among populations and the acclimation capacity and phenotypic plasticity
to OA differ greatly both within and between populations of same species and that understanding such variations
will be essential for predicting the impacts of ocean acidification.
INTRODUCTION
Current global climate change is occurring at an
unprecedented rate due to increasing anthropogenic carbon
dioxide (CO2) emissions. Consequently, significant
alterations in oceanic conditions are predicted over the next
century, including a reduction in surface water pH (termed
Ocean Acidification ‘OA’). OA represents an
unprecedented hazard to marine ecosystems, having the
potential to affect organismal physiological performance,
fitness and ultimately impact ecosystem biodiversity and
function. Physiological tolerances to environmental factors
limit the distribution of marine species. Geographically
widespread species may distribute across biogeographic
breaks. These transitional areas serve as limits to dispersal
or adaptation affecting patterns of abundance and variation
in phenotypic traits within species and populations. The
coast of Chile shows a transitional biogeographic break, in
an area around 30-32°S that limits an abrupt discontinuity
in upwelling regimes. Upwelling regions present surface
waters with low temperature and pH levels and are
predicted to be one of the environment most strongly
affected by OA [1]. The study of marine benthic organisms
that inhabit this area is a useful tool to predict organismal
and populational responses to OA. However, only few
investigations have considered the change in phenotypic
plasticity among geographic populations in marine
organisms to OA expected scenarios [2]. Consequently, our
study explores the following questions;
Vary the effects of OA among populations in same species
across a biogeographic break? Are there populational
differences in physiological traits at intraspecific level that
move beyond a biogeographic break? Is there a similar
pattern of pH acclimation capacity between populations ?
Is the variability of seawater carbonate system correlated
with phenotypic plasticity of populations?
MATERIAL AND METHODS
Individuals of populations of the marine mollusc Scurria
araucana were collected randomly by hand at low tides
from intertidal rocky shores situated at three localities
along the biogeographic break zone in Chilean coast
(Fig.1); Huasco (27° 59' S, 71°18′W), Talcaruca (30° 29'
89
(2016)
Pacific Ocean
Atlantic Ocean
Huasco
Talcaruca
Los Molles
Fig. 1. Location of the study sites along coast of Chile.
RESULTS AND DISCUSSION
Seawater carbonate chemistry parameters varied
significantly among study sites. The highest variability in
these parameter levels was recorded in Talcaruca site,
which is situated front of the upwelling center (Table 1).
S. araucana showed differences in scaling relationship
between populations. Furthermore, significant differences
(p<0.01) were observed for buoyancy weight and
metabolic rate between individuals collected from Huasco,
Talcaruca and Los Molles (Fig. 2) Regarding to the CO2
exposure experiment, population of S. araucana from Los
Molles showed the lowest metabolic rate at the 400 ppm
CO2 treatment
(i.e. control) while organisms from
Talcaruca showed the lowest metabolic rate was recorded
at the 1300 ppm CO2 treatment.
Results indicated that S. araucana populations from three
selected study sites, show different acclimation ability to
OA. In future OA conditions, population from Talcaruca
would be less impacted than those populations from
Huasco and Los Molles, as this site is characterized by a
higher environmental heterogeneity in terms of seawater
carbonate chemistry. Therefore, the environmental
heterogeneity would help the presence of phenotypic
plasticity, leading to a differential response of populations
to future OA scenarios.
Table 1. Summary of salinity, temperature and seawater
carbonate chemistry variables measured in three study
locations. Temperature is expressed as ºC. Total alkalinity
(TA), TIC, HCO3−, CO32− and CO2 are expressed in mM
kg-1 seawater. pCO2 is expressed as μatm.
Parameter
Salinity
Temperature
pHNBS
TA
DIC
CO3
pCO2
Ωcalcite
Ωaragonite
Metabolic rate (mgO2 L h-1g-1)
S, 71°41′ W ) and Los Molles (32° 24’ S, 71º 50´W). This
sites show differential variability in physical-chemical
characteristics of coastal waters. Organisms were
transported to the laboratory and maintained at constant
temperature (14 ºC) and salinity (33 ppt), aerated seawater,
for one week before experimental exposure and
measurements of oxygen consumption and cardiac activity
(Heart rate, HR). Furthermore, animals were characterized
in terms of shell length (mm), total buoyancy weight (g),
dry tissue and shell weight (g). After acclimation period a
subsample of individuals of S. araucana from Talcaruca
and Los Molles were exposed for Oxygen consumption
and buoyancy weights were measured at day 10, 20 and 30.
Oxygen consumption (O2 mg x L−1) was measured as
described by [2]. Cardiac activity was used as a measure of
variability in seawater carbonate parameters on
physiological performance. Cardiac activity was measured
as described in [3] and expressed as heart rate x min-1 .
Huasco
Talcaruca
Los Molles
34.50 ± 0.10
32.70 ± 0.34
32.06 ± 0.32
14.70
8.15 ± 0.12
2235.2 ± 20.0
2045.1 ± 82.6
155.3± 60.2
420.1 ± 122.1
3.96 ± 1.02
2.55 ± 0.82
13.80
7.90 ± 0.41
2279.6 ± 64.3
2158.2± 146.4
103.2 ± 39.6
879.9 ± 332.3
2.51 ± 0.95
1.60 ± 0.59
14.24
8.11 ± 0.14
2225.9 ± 30.2
2037.9 ± 145.1
144.3 ± 34.8
454.1 ± 176.0
3.51 ± 1.59
2.24 ± 1.01
140
B
120
B
100
80
60
A
0
Huasco
Talcaruca Los Mollles
Latitude
Fig. 2. Metabolic rate values in S. araucana populations
from three selected study sites. Capital letters represent
significant differences (p<0.01) between populations.
Mean±SE.
ACKNOWLODGEMENTS
This work was funded by FONDECYT Grant No. 1140092 and
by the Millennium Nucleus Center for the Study of Multiple
drivers on Marine Socio-Ecological Systems (MUSELS,
MINECON C120086).
REFERENCES
1 - Gruber N et al., 2012 Rapid Progression of Ocean
Acidification in the California Current System, Science.
2 - Lardies MA et al., 2014. Differential response to ocean
acidification in physiological traits of Concholepas
concholepas populations. J. Sea Res, 90, 127–134.
3 - Gaitán-Espitia JD et al.,2014 Geographic variation in
thermal physiological performance of the intertidal crab
Petrolisthes violaceus along a latitudinal gradient. J. Exp.
Biol.,
217,
4379–4386.
90
(2016)
What is the role of the main inlet of Ria Formosa coastal lagoon in the
exchanges with the Ocean? A seasonal approach
Alexandra Rosa1, Alexandra Cravo1 & José Jacob1
1
CIMA-Centro de Investigação Marinha e Ambiental, FCT, Universidade do Algarve, Campus de Gambelas, 8005-139 Faro,
Portugal
ABSTRACT
Ria Formosa lagoon is a multi-inlet barrier system on the South coast of Portugal, formed by 5 islands, 2
peninsulas and 6 inlets in permanent contact with the Atlantic Ocean. These inlets can be divided into three
distinct regions: east, central and west sectors, from where the Faro-Olhão inlet is the most important one in
terms of exchanges. The aim of this research addresses the need for improved understanding of the role of this
inlet in terms of seasonal mass exchanges (water, nutrients, chlorophyll a and suspended solids) with the
adjacent ocean, given the high biological productivity importance of this system. To perform this study, in situ
measurements and water samples along a selected cross section were accomplished hourly during 6 complete
tidal cycles, covering the different annual seasons. Results reveal that the mass balances depend on a multi-scale
temporal fluctuations – tidal and seasonal. The main driving mechanisms explaining the exchanges variability
are the biological productivity of the waters and physical processes acting on the adjacent coast. Seasonally, the
higher growth of phytoplankton reflected in the chlorophyll a (used as a proxy) was recorded during the springsummer campaigns when there was a relevant export of nutrients, chlorophyll and suspended solids, contributing
to fertilize the coast. The occurrence of upwelling, in autumn condition of spring tide revealed to be a key
process, having a main role in this coast, driving the fertilization of the Ria Formosa. The high amounts of mass
imported into this system contribute to replenish the nutrients and further increase its biological productivity.
1. INTRODUCTION
2. MATERIAL AND METHODS
Coastal lagoons are water bodies located in transition zones
between land and ocean. Consequently, these are very
dynamic and changeable over time. This study is focused
on Ria Formosa system (south coast of Portugal), with 55
km long (E-W direction) and a maximum width of 6 km
(N-S direction). It is considered a productive shallow
lagoon, having 5 islands, 2 peninsulas and 6 inlets.
Hydrodynamically, those are divided into three different
sectors: east, central and west. The last sector includes
Ancão, Faro-Olhão and Armona inlets and is the most
important one in terms of water exchanges [1]. It is
responsible for ~90% of the lagoon’s total tidal prism, from
which 59-71% circulates through Faro-Olhão, 25-37%
through Armona and <6% cross Ancão [1][2]. The tides are
semi-diurnal, controlling not only the water renewal (about
50-75% in each tidal cycle; [3]), but also the mass
exchanges. The main goal of this study at the main inlet of
Ria Formosa – Faro-Olhão was: a) to assess the seasonal
and tidal variability of mass exchanges of nutrients,
chlorophyll a and suspended matter with the Atlantic
Ocean, b) understand the interplay of the forcing
mechanisms between both environments and c) evaluate
the impact of the magnitude of those exchanges upon the
phytoplankton activity.
2.1 Field Site, Campaigns and Methods
Six seasonal oceanographic campaigns (2012-2013) were
conducted in a cross section at the Faro-Olhão inlet under
spring and neap tides in Spring and Autumn conditions.
Field measurements and samples were carried out over
complete semi-diurnal tidal cycles (~ 12.5 h) at three
different stations of the cross-section (center of the inlet
and both margins). In situ temperature, salinity, pH and
dissolved oxygen were measured using a multi-parametric
probe YSI (Model 6820). Water samples for the
determination of nutrients, suspended solids and
chlorophyll a were collected hourly at three depths:
Surface, Secchi disk extinction depth and bottom, using a 5
l Niskin bottle. To quantify the mass exchanges, the flow
velocity was measured using an ADP (ADP Bottom Track,
Sontek) and the variation of the sea level height was
measured by 2 pressure transducers (Level TROLL). The
nutrients and chlorophyll a concentrations were determined
spectrophotometrically, following the methods described in
literature [4 and 5, respectively]. The suspended solids
were determined using a gravimetric method [6]. The
organic and inorganic fractions of those were also
determined after combustion of the filters at 450ºC during
4h. The flood, ebb and residual tidal prisms were
determined by integrating the hourly flow rate along the
tidal cycle. Mass exchanges were obtained multiplying the
91
(2016)
flow rate by the average concentrations for the section,
then integrating them along the sampling tidal period.
RESULTS AND DISCUSSION
Data revealed that the variability of parameters depend on
multiscale temporal fluctuations, i.e. between fortnightly
tidal conditions, Spring tide (ST) vs. Neap tide (NT), and
seasonal conditions. Between tides the highest variability
along the tidal cycles was observed during spring tides.
However, the main processes responsible for the results
were both the higher biological activity in Spring and
Summer and the occurrence of upwelling in Autumn. This
is a frequent process in the south coast of Portugal [7] if the
wind is favourable (with strong west component). After
these periods nutrients and chlorophyll a are supplied to the
coast but can also have influence inside the Ria Formosa,
since an import of these compounds from the adjacent
ocean by tidal influence was recorded. The reflection of
these observations is mirrored in the estimates of mass
exchange through Faro-Olhão inlet (Table 1).
Table 1. Net tidal prisms and mass exchanges of nutrients,
Chlorophyll a (Chl a) and Suspended Solids (SS) during 6
campaigns. Positive values correspond to import into the
Ria Formosa and negative values to export to the Ocean.
Net Mass exchanges (kg)
NH4
Chl a
SS
-276
-203
25.8
-193 ton
-73.8
-15.6
12.8
28 ton
-113
-380
-14.3
-110 ton
21.7
43.5
-97.6
2.4
~ 4 ton
92.7
1290
-253
4.2
-210 ton
4.8
851
-158
Campaign
3
Prism (m )
SiO4
Spring - ST
5.42E+05
-1050
7.7
Spring - NT
7.31E+06
294
126
Summer - ST 7.93E+05
-1040
-306
Autumn - NT -3.11E+05
-68.1
Autumn - ST -1.45E+06
371
Winter - ST
-642
-1.97E+05
PO4
NO3
-13.8 -252 ton
In Spring and Summer campaigns this inlet behaved as a
flood inlet. In opposition, in Autumn and Winter
campaigns it behaved as an ebb inlet. Consequently, it
showed a changeable behaviour and it cannot be stated that
this is a consistent flood inlet, as is sometimes reported in
literature [3]. In Spring - ST, chlorophyll a reached the
maximum concentration range (0.7-2.5 µg/L), and it
corresponded to a total import of ~ 26 kg, along with ~8 kg
of phosphate. The remaining nutrients (specifically silicate
~ 1 ton) and SS (193 tons) were exported. As the organic
fraction of the SS represents ~55% it means that by that
time ~100 ton of organic matter contributed to fertilize the
adjacent coast. In Spring - NT there was a general import
of mass, except for ammonium and nitrate that were
exported from the Ria Formosa. In Summer campaign,
there was a maximum export of nutrients. By that time ~14
kg of chlorophyll a was exported, reflecting its higher
productivity inside the lagoon than in the coast. As ~ 26%
of the SS corresponded to organic fraction, it represented ~
30 ton to further fertilize the adjacent sea. In Autumn – NT,
there was an import of phosphate, nitrate, chlorophyll a and
SS, whereas in Autumn - ST, there was an evident increase
of nutrients import (~ 90 kg phosphate and > 1 ton nitrate)
except for ammonium and SS. This can be explained
because this last campaign occurred during an upwelling
event, with no wind relaxation, not promoting the
phytoplankton growth. In the winter, despite the export of
water, chlorophyll a, ammonium, silicate, and suspended
solids (majorly inorganic ~ 82%), there was an import of
phosphate and nitrate from the coast which, like in
Autumn, would contribute to replenish the nutrients and
increase the productivity of the Ria Formosa.
In summary, at the main inlet of this system – Faro-Olhão,
regardless the tide and the net prism of the tide, the higher
productivity of the Ria Formosa during Spring and
Summer seasons contributes to fertilize the coastal zone,
particularly in ammonium and nitrate (that showed to be a
source to the coast). In Autumn and Winter seasons, as the
nutrients will decrease inside the lagoon by continuous
consumption from phytoplankton, even if the inlet behaved
as an ebb one, phosphate and nitrate were imported. This
shows a deficiency of those nutrients comparatively with
their higher availability on the coast. The magnification of
the import of nutrients in Autumn - ST was also coincident
with the upwelling event, the driving mechanism supplying
the nutrients into the lagoon and contributing to further
increase its biological productivity.
ACKNOWLEDGEMENTS
The authors are grateful to all the team members for their
support during the campaigns. This work was financially
supported by FCT (Portuguese Foundation for Science and
Technology)
under
the
project
ref:
“PTDC/MAR/114217/2009-COALA“.
REFERENCES
1 - Pacheco, A., Ferreira, Ó., Williams, J. J., Garel, E.,
Vila-Concejo, A., & Dias, J. A., 2010. Hydrodynamics and
equilibrium of a multiple-inlet system. Marine Geology,
274(1-4), 32–42;
2 - Jacob, J., Cardeira, S., Rodrigues, M., Bruneau, N.,
Azevedo, A., Fortunato,A. B., Rosa, M., Cravo, A., 2013.
Experimental and numerical study of the hydrodynamics of
the western sector of Ria Formosa. Journal of Coastal
Research, 65, 2011–2016;
3 - Newton, A., & Mudge, S. M., 2003. Temperature and
salinity regimes in a shallow, mesotidal lagoon, the Ria
Formosa, Portugal. Estuarine, Coastal and Shelf Science,
57(1-2), 73–85;
4 - Grasshoff, K., Erkhardt, M. and Kremling, K. 1983.
Methods of Seawater Analysis. Verlag Chemie, New York,
419 pp;
5 - Lorenzen Carl J., 1967. Vertical distribution of
chlorophyll and phaeo-pigments: Baja California. Deep Sea
Research and Oceanographic Abstracts, 14 (6): 735- 745;
6 - APHA, A., 2002. WPCF, Standard Methods for the
Examination of Water and Wastewater. Amer. Public
Health Assoc., Washington, DC;
7 - Relvas, P., Barton, E.D., 2002. Mesoescale patterns in
the Cape São Vicente (Iberian Peninsula) upwelling region.
J. Geophys. Res. 107 (C10), 3164 (28 (1 23)).
92
(2016)
Efectos de la carbamazepina sobre la termorresistencia de Artemia
parthenogenetica
Raquel Samper 1, Ivan Morant 1, Deborah Aurora Perini2, Inmaculada Varó3 & Amparo
Torreblanca1
1
Departamento de Biología Funcional y Antropología Física. Universitat de València (España)
Universidad de Siena (Italia)
3
Instituto de Acuicultura Torre de la Sal. CSIC (España)
2
RESUMEN
En las última décadas se ha realizado un gran esfuerzo para determinar los niveles de las denominadas sustancias
emergentes que llegan a los diferentes compartimentos ambientales. Entre estas sustancias se encuentran los
fármacos de uso humano o veterinario. La carbamazepina (CBZ) es un medicamento antiepiléptico de uso
clínico y se ha detectado su presencia en el medio acuático [1]. Aunque recientemente se han llevado a cabo
estudios que contribuyen a conocer los efectos que este medicamento ejerce sobre los crustáceos [2], todavía
quedan muchos aspectos por explorar. Artemia constituye un genero de crustáceos que presenta una serie de
adaptaciones fisiológicas que le permite vivir en medios hipersalinos (ej. las salinas, lagos y lagunas saladas),
cuerpos de agua poco favorables para la vida en cuanto a la salinidad, la temperatura y el oxígeno disuelto [3].
Artemia parthenogenetica es una de las dos especies autóctonas que conviven en las salinas del Parque Natural
de las Lagunas de La Mata y Torrevieja (Alicante) [4]. El objetivo del presente trabajo es evaluar si la
exposición a CBZ puede alterar la termorresistencia natural y la termorresistencia adquirida mediante choque
térmico en la especie partenogenética autóctona de la Mata Artemia parthenogenetica. Para ello, quistes
recolectados en la Salinas de la Mata se eclosionaron según el procedimiento descrito por Sarabia et al [4]. Tras
la eclosión los nauplios fueron recogidos y mantenidos en condiciones estándar de cultivo (agua de mar y 25ºC)
durante una semana. A continuación los individuos fueron expuestos a 1 µg/L y 100 µg/L de CBZ durante 7
días. Se encontró una mayor supervivencia en los individuos expuestos a 1 µg/L de CBZ respecto al grupo
control. Tras el período de exposición a la CBZ, y en medio carente de fármaco, se determinó la
termorresistencia natural siguiendo el procedimiento descrito por Clegg et al [5], observándose que un choque
térmico de 42ºC durante 45 minutos provocaba una menor mortalidad (57%) en el grupo expuesto previamente a
100 µg/L de CBZ respecto a los otros dos grupos experimentales (> 95%). Sin embargo, cuando los animales
fueron sometidos a un choque térmico de 39ºC durante 45 minutos en el día previo al choque térmico de 42ºC
(termorresistencia adquirida), no se encontraron diferencias debidas a la exposición a CBZ. Se están realizando
más estudios para establecer la relación concentración-respuesta del efecto de la CBZ sobre la termorresistencia
y evaluar los efectos sobre otros parámetros fisiológicos y bioquímicos.
REFERENCIAS
1 - Boxall ABA, Keller VDJ, Straub JO, Monteiro SC,
Fussell R, & Williams, RJ, 2014. Exploiting monitoring
data in environmental exposure modelling and risk
assessment of pharmaceuticals. Environ Int, 73:176-185.
2 - Nieto E, Hampel M, González-Ortegón E, Drake P&
Blasco J, 2016. Influence of temperature on toxicity of
single pharmaceuticals and mixtures, in the crustacean A.
Desmarestii. J. Hazard. Mater.,313:159-169.
3 - Amat F, 1985. Biologia de Artemia. Informes Técnicos
Instituto Investigaciones Pesqueras., 127: 3-60.
4 – Sarabia R, Del Ramo J, Varó I, Díaz-Mayans, J &
Torreblanca, A, 2008. Sublethal zinc exposure has a
detrimental effect on reproductive performance but not on
the cyst hatching success of Artemia parthenogenetica. Sci
Total Environ, 398(1–3):48-52.
5 – Clegg JS, Jackson SA, Van Hoa N & Sorgeloos P,
2000. Thermal resistance, developmental rate and heat
shock proteins in Artemia franciscana, from San Francisco
Bay and southern Vietnam . J. Exp. Mar. Biol. Ecol., 252:
85–96
93
(2016)
Trace metal accumulation in marine macrophytes: Hotspots of coastal
contamination worldwide
David Sánchez-Quiles1, Núria Marbà1, Antonio Tovar-Sánchez1,2
1
Department of Global Change Research. Mediterranean Institute for Advanced Studies, IMEDEA (CSIC-UIB), Miguel
Marqués 21, 07190 Esporles, Balearic Island, Spain.
2
Department of Ecology and Coastal Management. Andalusian Institute for Marine Science, ICMAN (CSIC). Campus
Universitario Río San Pedro, 11510 Puerto Real, Cádiz. Spain.
ABSTRACT
This study compiles, from 155 peer review research articles, almost 23,000 estimates of trace metals (As, Cd, Co,
Cr, Cu, Fe, Hg, Mn, Ni, Pb and Zn) contents in natural populations of marine macrophytes (seagrasses,
chlorophytae, phaeophytae and rhodophytae) worldwide. The objective was to explore the distribution of these
metals, to examine its trends of accumulation and to identify hotspots of coastal pollution. Our results point out
phaeophytae as the group with the maximum accumulation capacity and tolerance to elevated concentrations of
metals, and indicate that, despite of the geographic differences and local and regional influences, exist a similar
atomic ratio of trace metals accumulation in the four groups of marine macrophytes regardless the species. The
mapping of geographic distribution of metal accumulation in marine macrophytes identifies some coastal areas as
hotspots of trace metal contamination. This work aims to provide a reference for futures studies.
INTRODUCTION
MATERIALS AND METHODS
Studies of trace metals concentrations in marine
macrophytes are essential to elucidate their role in the
ecology and oceans. While global assessments of
concentrations and requirements of metals in marine
organisms have been examined for phytoplankton and for
seagrass leaves [1,2], information about the most
representative macrophytes is almost nonexistent.
In marine macrophytes Co, Cu, Fe, Mn, Ni and Zn act as
micronutrients working as cofactors in several enzymes
and vitamins, and in several metabolic paths [3,4].
However, As, Cd, Cr, Hg or Pb do not have a essential
biological role in plant’s metabolism and even in low
concentrations have negative effects on plant growth [5–7].
Since metal accumulation in marine macrophytes depends
strongly on the specie, location, and season [8], a global
overview of metal composition might help to better
understanding of biological requirements and tolerance
under different environments conditions.
Here, we have assessed the variability in accumulation of
major trace metals in seagrasses and in the three groups of
macroalgae (phaeophytae, chlorophytae and rhodophytae)
and we have identified some hotspots of coastal
contamination by compiling reported estimates of trace
metal concentrations in marine macrophytes worldwide.
Results from this meta-analysis will be useful to compare
future results from studies on metal composition in
macrophytes and to have a baseline that allows assess and
detect natural or anthropogenic environmental changes.
The data set includes concentration of eleven trace metals
(As, Cd, Co, Cr, Cu, Fe, Hg, Mn, Ni, Pb and Zn) in marine
macrophytes worldwide. From each study we recovered
latitude and longitude, macrophyte species, macrophyte
group, concentration of each metal and year of publication.
Due to the great number of observations collected we have
used robust statistics estimators to characterize the
populations: the 0.2-truncated mean, used as robust
measure of the central tendency, and the 0.2-Winsorized
Standard Deviation, used as a robust estimation of the
variability. As the robust, the resampling methods provide
consistent results. These are based on repeated sampling
within the same sample. Here we used one of the most
implemented resampling techniques: the bootstrap, in order
to constructing hypothesis tests through the estimation of
intervals to compare differences between two independent
populations. For each analysis we performed 1,000
resamples of each sample.
RESULTS AND DISCUSSION
The data set includes values of trace metal concentrations
in marine macrophytes reported in 155 peer review papers
published between 1956 and 2014. We compiled a total of
22,969 estimates of concentration for different metals: As
(2.4 %), Cd (13.5 %), Co (4.6 %), Cr (6.1 %), Cu (14.0 %),
Fe (10.1 %), Hg (4.3 %), Mn (9.2 %), Ni (8.5 %), Pb (12.4
%) and Zn (14.7 %) in seagrasses (34.9 %), chlorophytae
94
(2016)
(25.9 %), phaeophytae (22.5 %) and rhodophytae (16.7 %)
with a widely geographical distribution.
Our results reveals the different capacity of the marine
macrophytes to accumulate metals in their tissues with a
significance level of 0.05: We have found that 1)
phaeophytae present significantly higher concentrations
than seagrasses, chlorophytae and rhodophytae of As, Cd,
Co and Zn; 2) Chlorophytae shows the highest
concentrations of Fe and Mn; 3) Seagrasses present the
highest concentration of Cu and the lowest of Hg and Mn;
and 4) Rhodophytae have significantly lower
concentrations in 5 of the 11 trace metals studied, than the
other plants: i.e. Cr, Cu, Fe, Pb and Zn. Moreover, the
atomic ratio of metals concentrations in each group show
similar pattern of accumulation suggesting that the ability
to accumulate trace metal is equal in all macrophytes,
regardless the different species and locations.
Due to marine macrophytes are good tools to control metal
concentrations, we have identified some hotspots of trace
metals in coastal areas worldwide (Fig. 1). Some of these
areas are clearly influenced by human activities that
increase the amount of metals discharged in the marine
ecosystem. That is the case of the coasts of the
Mediterranean and the Black Sea, where we have found
many hotspots of trace metals pollutions (i.e. Cd, Cr, Cu,
Fe, Hg, Ni, Pb and Zn). The gulfs region of South Australia
(Spencer Gulf and Gulf St Vicent) is another example of
metal contamination by anthropogenic activities. We have
found in this region, elevated concentrations of As, Cd, Pb
and Zn.
The Bay of Bengal, in the Indian Sea, is another hotspot of
trace metal contamination due to the elevated concentration
of metals found in marine macrophytes. Concretely, in the
coasts of Andaman Island we have found elevated
concentrations of Cr, Cu and Mn.
The particular geography of Chile and the upwelling that
occurs in its coast could increase the concentrations of
trace metals in marine macrophytes. We have found along
the coast of Chile hotspots of Cr, Cu, Fe, Ni and Zn.
It is remarkable that, although the Antarctic is far from
human influence, its coast presents elevated concentrations
of As, and Co in plants collected from King George Island.
This could be due to natural processes like the volcanic
composition of the rocks near the sampling place. This
natural origin of As contrast with other hotspots found in
other bays from England, France and China. It is also
notable that the Gulf of Mexico does not present any
hotspot instead of its strong anthropogenic pressure.
The extensive database of metal concentration in marine
macrophytes compiled in the present work could help in
future monitory programs of the coastal pollution and
provide a reference for futures studies.
Fig. 1. Worldwide hotspots of trace metals.
REFERENCES
1. Govers LL, Lamers LPM, Bouma TJ, Eygensteyn J, de
Brouwer JHF, Hendriks AJ, Huijbers CM, van Katwijk
MM. Seagrasses as indicators for coastal trace metal
pollution: A global meta-analysis serving as a benchmark,
and a Caribbean case study. Environ. Pollut. 2014
Dec;195:210–217.
2. Twining BS, Baines SB. The trace metal composition of
marine phytoplankton. Annu. Rev. Mar. Sci. 2013;5:191–
215.
3. Fageria NK, Baligar VC, Clark RB. Micronutrients in
Crop Production. In: Donald L. Sparks, editor. Adv. Agron.
[Internet]. Academic Press; 2002 [cited 2014 Feb 25]. p.
185–268.
4. Marschner H. 9 - Functions of Mineral Nutrients:
Micronutrients. In: Marschner H, editor. Miner. Nutr. High.
Plants Second Ed. [Internet]. London: Academic Press;
1995 [cited 2014 Apr 26]. p. 313–404.
5. Lamai C, Kruatrachue M, Pokethitiyook P, Upatham
ES, Soonthornsarathool V. Toxicity and Accumulation of
Lead and Cadmium in the Filamentous Green Alga
Cladophora fracta (O.F. Muller ex Vahl) Kutzing: A
Laboratory Study. ScienceAsia. 2005;31:121–127.
6. Stewart JG. Effects of lead on the growth of four
species of red algae. Phycologia. 1977 Mar;16:31–36.
7. Strömgren T. The effect of lead, cadmium, and mercury
on the increase in length of five intertidal fucales. J. Exp.
Mar. Biol. Ecol. 1980;43:107–119.
8. Hou X, Yan X. Study on the concentration and seasonal
variation of inorganic elements in 35 species of marine
algae. Sci. Total Environ. 1998 Oct;222:141–156.
95
(2016)
Efecto de la acidificación oceánica en la química del hierro y su
interacción con la materia orgánica
J. Magdalena Santana-Casiano
Instituto de Oceanografía y Cambio Global. Universidad de Las Palmas de Gran Canaria
RESUMEN
El efecto del incremento del CO2 de origen antropogénico en el sistema de carbonatos en el océano y la
consecuente disminución del pH, es una de las líneas de investigación oceanográfica más relevantes en el
contexto del cambio global. La variación de pH afecta directamente a la especiación de metales traza
esenciales para el desarrollo de los organismos fitoplanctónicos y, por lo tanto, a su comportamiento
biogeoquímico en el medio marino, debido a los cambios que se producen en los equilibrios ácido-base,
oxidación-reducción, solubilidad-precipitación y complejación. La variación de pH también va a afectar
al fitoplancton marino desde una perspectiva química, modificando el grado de protonación de los grupos
superficiales de la pared celular y afectando a la especiación de los productos orgánicos excretados. Estos
cambios, tanto en la especiación del metal traza, como en los grupos funcionales de los compuestos
orgánicos excretados por los organismos y por el propio cambio en la pared celular van a repercutir en el
comportamiento del metal y en su interacción con el fitoplancton. Sin embargo, se tiene un gran
desconocimiento del efecto que puede tener un cambio de pH en la interacción del hierro con compuestos
orgánicos excretados por el fitoplancton, como los compuestos polifenólicos, aminoácidos, polisacáridos
y sideróforos en particular, y cómo afectará esto a la biodisponibilidad de hierro en el medio marino.
INTRODUCCIÓN
Aunque existen numerosas referencias sobre la
importancia de los metales traza, sus ciclos y aportes
en el medio marino, debido a proyectos
internacionales como GEOTRACES, gran cantidad
de información sobre el CO2 en los océanos,
potenciada por proyectos como CARBOCHANGE,
CARBOOCEAN o ATLANTOOS, y sobre los
estudios de acidificación oceánica y efecto en las
comunidades biológicas, iniciados a partir de los
proyectos EPOCA y BIOACID, la investigación
sobre la influencia del aumento del CO2, en cuanto a
acidificación y calentamiento, y la investigación
sobre la biogeoquímica de los metales traza,
continúan
tratándose
como
disciplinas
independientes.
El pH y la temperatura son dos variables que
controlan los procesos químicos y biológicos en el
medio marino y, por lo tanto, las investigaciones
relacionadas con la biogeoquímica de los metales
traza, la acidificación océanica y el calentamiento de
las aguas superficiales deberían estar ligadas. Las
publicaciones relacionadas con estudios en el océano
sobre las múltiples interacciones entre el pH, la
temperatura y los metales traza son escasas, pero dan
cuenta de la complejidad y el significado que tendría
el realizar este tipo de estudios [1,2] promovidos
también por el programa internacional IMBER
(Integrated Marine Biogeochemistry and Ecosystem
Research). Dentro de las estrategias del programa
IMBER, en el tema correspondiente a la Sensibilidad
al Cambio Global, se plantea el conocer cuáles son
las
respuestas
que
presentan
los
ciclos
biogeoquímicos claves, ecosistemas
y sus
interacciones debidas al cambio global y se propone
el estudio del efecto del incremento de CO2
atmosférico y cambio de pH en los ciclos
biogeoquímicos y sus interacciones. Y en concreto, el
efecto que los cambios de pH producen en la
biodisponibilidad y especiación de metales esenciales
para el desarrollo del fitoplancton marino. En este
trabajo se realiza una revisión de los estudios
realizados por el grupo QUIMA de la ULPGC para
entender el efecto que el calentamiento y la
acidificación oceánica tienen en el ciclo del hierro
(Fig. 1) en el medio marino.
EVOLUCIÓN DE LOS ESTUDIOS
PLANTEADOS. RESULTADOS Y DISCUSIÓN
Efecto del calentamiento y acidificación oceánica
en la especiación y cinética de oxidación del Fe(II)
en condiciones oligotróficas y eutróficas [3]. El
efecto de los nutrientes (nitratos, fosfatos y silicatos)
en la cinética de oxidación de Fe(II) a
concentraciones nanomolares se ha evaluado en
function del pH (7.2-8.2), la temperatura (5-35 ºC) y
la salinidad (10-37.09) observándose que la constante
que define la velocidad de oxidación del Fe(II) (kapp)
es mayor en presencia de nutrientes produciéndose
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(2016)
las mayores variaciones a temperaturas más altas, en
todo el rango de pH considerado.
La disminución de pH reduce la velocidad de
oxidación del Fe(II). La especiación de Fe(II) que
controla el proceso está constituida por Fe2+, FeOH+,
Fe(OH)2,
FeCO3(OH)-,
FeCO3,
Fe(CO3)22-,
+
FeH3SiO3 , FePO4 .
Interacción de Fe(II) con otros metales esenciales
como el Cu [4,5]. La presencia de Cu(I) y Cu(II) en
el medio produce una aceleración en la oxidación de
Fe(II) que se explica por un acoplamiento redox entre
el Fe y el Cu, por una competición por especies
inorgánicas (hidroxilos y carbonatos) y por la
formación de precipitados Fe-Cu, jugando los
intermedios de oxidación de oxígeno (O2· y H2O2) un
papel importante en este proceso.
Producción de exudados por el fitoplancton [6,7].
P. tricornutum y D. tertiolecta excretan al medio
polifenoles cuyas concentraciones y composición
depende de la cantidad de Fe y Cu presente en el
medio. Esta respuesta esta relacionada con las
necesidades del fitoplancton por estos metales.
Efecto de los exudados de fitoplacton en la cinética
de oxidación del Fe(II) [8,9]. Los estudios realizado
en presencia de exudados de P. tricornutum y D.
tertiolecta en diferentes condiciones de pH,
temperatura y salinidad demuestran que kapp
disminuye con el pH, T y S, por lo que la presencia
de estos compuestos orgánicos favorecen la presencia
de Fe(II) en el medio.
Efecto de los compuestos orgánicos en el ciclo
redox del Fe [10,11]. Los compuestos polifenólicos
como el catecol producen una reducción del Fe(III) a
Fe(II), lo que se ve favorecido por una disminución
en el pH. Esto se explica por la competición que se
produce entre el Mg(II) y el Fe(III) por el radical
semiquinona implicado en el proceso de reducción de
Fe(III). Este mismo comportamiento se observa para
dos polifenoles excretados por P. tricornutum y D.
tertiolecta, el catequin y el ácido sinápico. El estudio
se realiza bajo diferentes escenarios de acidificación
oceánica.
Modelización del comportamiento del hierro en
diferentes escenarios de acidificación oceánica
Los estudios en mesocosmos [12] y en zonas de
emission hidrotermal submarina de CO2, como los
realizados en la zona del volcán submarino de El
Hierro [13,14], nos permiten ampliar nuestro
conocimiento sobre el comportamiento del hierro en
condiciones de acidificación, convirtiéndose el area
en un laboratorio de experimentación natural en el
que se combinan multiples factores.
Fig. 1. Ciclo redox del hierro en el medio marino
AGRADECIMIENTOS
Estos estudios han sido financiados por los proyectos
ECOFEMA (CTM2010-19517-mar) y EACFe
(CTM2014-52342-P) del Ministerio de Economía y
Competitividad. Los estudios en el volcán submarino
de El Hierro han sido posibles por CARBOCHANGE
(264879)
VULCANO
(CTM2012-36317)
y
VULCANA (IEO, 2015-2017).
Para entender, explicar y modelizar los procesos que
controlan el ciclo biogeoquímico del hierro, es
necesario combinar los resultados obtenidos a partir
de estudios de laboratorio, en los que se pueden
controlar y modificar las condiciones experimentales,
con los obtenidos en el medio marino.
REFERENCIAS
1 - Boyd PW & Ellwood MJ, 2010. The
biogeochemical cycle of iron in the ocean. Nat.
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8 - González AG et al., 2012. Effect of organic
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Phaeodactylum tricornutum and their effects on the
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12 - Breitbarth E. et al., 2010. Ocean acidification
affects iron speciation during a coastal seawater
mesocosm experiment. Biogeosciences 7:1065-1073.
13 - Santana-Casiano JM et al., 2013. The natural
ocean acidification and fertilization event caused by
the submarine eruption of El Hierro. Sci. Reports
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14 - Santana-Casiano JM et al., 2016. Significant
discharge of CO2 from hydrothermalism associated
with the submarine volcano of El Hierro Island. Sci.
Reports
6:25686
DOI:
10.1038/srep25686
98
(2016)
Behaviour of CeO2 nanoparticles and bulk and their toxicity in freshwater
and seawater microalgae
Marta Sendra1, Ignacio Moreno-Garrido1, Pilar Yeste2, José Manuel Gatica2 & Julián Blasco1.
1
Department of Ecology and Coastal Management, Institute of Marine Sciences of Andalusia (CSIC).Campus Río S.
Pedro.11510, Puerto Real, Cádiz, Spain.
2
Department of Inorganic Chemistry. Faculty of Sciences. University of Cádiz. Campus Río S. Pedro.11510, Puerto Real,
Cádiz, Spain
ABSTRACT
The production of manufactured nanomaterials has risen exponentially in recent years. Despite this growth,
information regarding the fate and behavior of nanoparticles (NPs) in freshwater and marine environment is still
limited. CeO2 NPs are used for polishing and decolorizing glass, opacifier in vitreous enamels and photochromic
glasses, heat-resistant alloy coatings; as a cracking catalyst, as a catalyst for automobile emission control, in
ceramic coatings, in phosphors, in cathodes, in capacitors, in semiconductors, in refractory oxides, gemstone
polishing.
The aim of the current study was to assess the toxicity of CeO2 in two species of microalgae from freshwater to
marine microalgae Three aspects were studied under laboratory conditions, I) CeO2 NPs and Bulk behavior in
different culture media used for bioassays, II) toxicity of different type of CeO2 NPs and bulk and III)
quantification of intracellular and extracellular NPs and Bulk CeO2. The experiment was performed for two
microalgae. Phaeodactylum tricornutum, and Chlamydomonas reinhardtii, seawater and freshwater species
respectively. Both studies were carried out under white light illumination.
CeO2 aggregation rate was increased by ionic strength. Moreover, the presence of algal cells affects the stability
of CeO2 suspensions, where heteroagglomeration process were present as the first mechanism of NP-cell
interaction. Both, nano and bulk CeO2 formed aggregates during incubation, but CeO2 NPs formed large
aggregates trapped almost completely between algae more than bulk CeO2 did, because NPs are more bioavailable
over bioassays.
INTRODUCTION
The growing use of manufactured nanomaterials in
consumer products have as result that engineered
nanoparticles (ENPs) are inevitably released into aquatic
systems, including oceans; coastal waters are the ultimate
sink form (ENPs) (1-3).
ENPs is raising questions as to whether nanosized
materials should be regulated differently to macroscopic
forms of the same compounds in terms of the risks they
pose both to human and ecosystem health (4-6).
The choice of CeO2 was reinforced by its wide potential
usage, particularly as an additive to diesel fuels where it
improves the combustion efficiency of engine carbon
deposits, reducing particulate emissions and improving fuel
efficiency (7). It has also been shown to be an effective
photocatalyst for water decomposition (8).
In microalgae, the cytotoxicity seems to be due to
membrane damage, impairment of the effective quantum
yield of PS II and cell cycle (9, 10).
The extent and type of damage depend on physicochemical
characteristics of CeO2 NPs (e.g., size, charge, crystalline
forms and coating) and environmental factors (e.g., ionic
strength (IS), pH and dissolved organic materials which
governs their bioavailability and reactivity)(11-13).
In this work, toxicity of different types of CeO2 NPs (NPs
in suspension water and powder) and bulk CeO2 on two
phytoplanktonic species (Paeodactylum tricornutm, from
marine, and Chlamydomonas reinhardtii, from freshwater
environment) were assessed. The joint effect of UV-A
radiation has been taken in account in order to improve the
knowledge of the mechanism involved in the CeO2 NPs
and bulk toxicity in environmental conditions.
MATERIAL AND METHODS
99
(2016)
Characterization of CeO2 NPs and bulk CeO2.
Textural characterization of samples was carried out by
measuring the absorption/desorption of N2 at 196 °C,
employing a Micromeritics ASAP 2010 automatic device.
Before measurements, samples were submitted to a surface
cleaning pre-treatment under high vacuum at 200 °C during
2 hours. The obtained isotherms were used to calculate the
specific surface area (SBET) as well as the micro- and mesoporosity features of studied samples.
Initial particle size of CeO2 NPs, as well as zeta potential
of CeO2 in both forms (NPs and bulk) were studied in
ultrapure water, freshwater and artificial marine water
through Dynamic Light Scattering (Zetasizernano
ZS90,Malvern, and its software version 7.10) at 1 mg·L-1.
CeO2 NPs dispersion was prepared in ultrapure water,
following the Standard protocol CEINT/NIST 1200-3 and
1200-4 (14, 15).
Particle size, shape and structure were confirmed by
Transmission Electron microscopy (TEM).
Agglomeration of CeO2 NPs was evaluated over time (0,
0.5, 1, 3, 6, 24 and 48 h) in different media (ultrapure
water, synthetic freshwater and synthetic marine water)(16,
17). Initial concentration in studied samples was 250 mg·L1
in freshwater and marine water, but for ultrapure water
concentrations of 250 mg·L-1 and 500 mg·L-1were
evaluated. Changes in agglomeration states were measured
by a Master Sizer 2000, Malvern with the software version
5.61.
Test organisms.
Two microalgal species were selected, one of them from
freshwater environment (Chlamydomonas reinhardtii P.A.
Dangeard (1888), CHLOROPHYCEAE) and another from
marine environment (Phaeodactylum tricornutum Bohlin
(1897), BACILLARIOPHYCEAE) both obtained from the
ICMAN Marine Microalgae Culture Collection (IMMCC).
Cells were grown in filtered (0.2 µm) freshwater culture
medium and F/2 marine medium lacking EDTA for two
weeks prior to the experiment (16, 18). Synthetic marine
water used was the Substitute Ocean Water D1141-75 from
ASTM (17).
Toxicity bioassays
Bioassays were carried out using CeO2 NPs or bulk CeO2.
Two light treatments were also applied: continuous visible
light (300 µE-2s-1) and the same light regime plus 6 h of
UV-A (0.20 mWcm-2, Multiple Ray Lamp) per day. The
intensity of UV-A was measured with a digital UVX
radiometer (UVP, Analytic Jena Company).
A series of screening assays were developed to determine
the potential for the CeO2 particles to EC 50% of growth
inhibition respect to the controls following OECD (1994)
Guidelines(19), effective quantum yield of photosynthetic
energy conversion in PSII in dark was measured by
fluorometry using a Phyto-PAM (Heinz Walz GmbH) (20),
reactive oxygen species (ROS) production were measured
by FACSCalibur Flow Cytometer (Becton-Dickinson®)
(21, 22) and membrane integrity was also quantified,
following the Propidium Iodide (PI) method (23) by the
flow cytometer.
RESULTS AND DISCUSSION.
Characterisation of the CeO2 particles (NPs and bulk) was
performed using a combination of techniques in order to
provide information on the particles, chemistry, surface
area, morphological shape, porosity and size distribution.
TEM analysis showed the significant difference in particle
size of the material studied. Despite all suspensions
showing considerable aggregation in different culture
media (freshwater and artificial marine water), the primary
particle size and aggregates size measured by DLS were
smaller for the nanoparticles into water suspension (105.8
and 196.4 nm) than power nanoparticles and bulk (731 and
4567 nm). Zeta potential of NPs were negative charged
between -18 and -56 mV.
The flow cytometric data indicate that nano CeO2 increases
ROS production and membrane permeability more than
bulk where growth of culture and effective quantum yield
were lower when culture were exposed to NPs.
Experiments under UV-A regime showed higher toxicity
because its photocatalytic properties, so this condition due
to be considered in bioassays which photorreactive
substances are used.
ACKNOWLEDGEMENTS
This work was supported by PE2011-RNM-7812 project
and Spanish National Research Plan (MINECO)
CTM2012-38720-C03-03. The ERF funds has been
supported this project. We would like to thank to Dr.
Catalina Fernández from IFAPA-CICEM El Toruño for her
support in the DLS measurements.
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Gonzalez L, Lison D, Kirsch-Volders M.
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Liu Y, Li S, Chen Z, Megharaj M, Naidu R.
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101
(2016)
Variabilidad estacional de la concentración de N2O en el Golfo de Cádiz:
Flujos agua-atmósfera
A. Sierra1, *, D. Jiménez-López1, T. Ortega1, R. Ponce1, M.J. Bellanco2, R. Sánchez-Leal2, A.
Gómez-Parra1 y J. Forja1
1
Dpto. Química-Física. CACYTMAR.Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz, Campus
Universitario Río San Pedro, 11510 – Puerto Real, Cádiz, Andalucía, España.
2
Instituto Español de Oceanografía. Centro Oceanográfico de Cádiz. Puerto Pesquero, Muelle de Levante s/n. Apdo. 2609. E11006, Cádiz (España).
*Correo del autor: [email protected]
RESUMEN
Se ha determinado la variabilidad estacional de N2O durante 2014 y 2015 y a lo largo de varias secciones
(Guadalquivir, Sancti Petri y Trafalgar) en el Golfo de Cádiz. Las concentraciones de N2O se han cuantificado
mediante un cromatógrafo de gases. Se ha observado un aumento de la concentración de este gas con la
profundidad, provocado por las propias características termodinámicas de la zona y por la remineralización
bentónica. La zona de estudio se comporta como una fuente de N2O a la atmósfera con emisiones globales de
0,21 y 0,17 Gg N2O año-1 en 2014 y 2015 respectivamente. Es en la sección de Guadalquivir dónde se producen
los flujos medios más elevados durante los dos años de estudio, probablemente relacionado con la mayor
producción primaria detectada en la zona.
INTRODUCCIÓN
El óxido nitroso a pesar de estar presente en menor
concentración que el CO2, tiene un efecto en el
calentamiento global 298 veces mayor [1] y además una
vez emitido, posee un tiempo de residencia en la atmósfera
de 114 años [2]. El N2O, se produce de forma natural por
diversas fuentes biológicas presentes en el suelo y en el
agua, aunque existen también fuentes antropogénicas,
como son el uso de fertilizantes, la quema de biomasa y
algunas actividades industriales [3].
El N2O generado en el medio marino aparece
principalmente por dos procesos microbianos, como son la
nitrificación y la desnitrificación. Ambos procesos pueden
tener lugar en los sedimentos, la columna de agua o en el
interior de partículas en suspensión. La producción
oceánica de N2O está principalmente relacionada con los
procesos de nitrificación, y sólo un 7% de este gas se
genera como intermedio de reacción en los procesos de
desnitrificación [4].
MATERIAL Y MÉTODOS
El área de estudio es la parte oriental del Golfo de Cádiz,
situado al suroeste de la península Ibérica. La
hidrodinámica en el Golfo de Cádiz se encuentra dominada
por el intercambio de masas de aguas que se produce en el
Estrecho de Gibraltar, entre el océano Atlántico y el mar
Mediterráneo. A la circulación general que tiene lugar en el
Golfo de Cádiz, hay que añadirle la entrada de aguas
continentales procedentes de diversos ríos, como el
Guadalquivir.
Las muestras se tomaron durante la realización de las
campañas STOCA de 2014 y 2015 a bordo de los buques
oceanográficos Ángeles Alvariño y Ramón Margalef. En
cada una de las campañas, se han llevado a cabo tres
secciones perpendiculares a costa, con diferentes estaciones
de muestreo en cada una y a determinadas profundidades.
Estas secciones se localizan en la desembocadura del
Guadalquivir, del caño de Sancti Petri y en el cabo de
Trafalgar.
Para el análisis de N2O, las muestras se tomaron por
duplicado en frascos Winkler de 250 mL, se fijaron con
HgCl2 para inhibir procesos microbiológicos, y se sellaron
con grasa Apiezon® para prevenir el intercambio gaseoso
con la atmósfera.
La concentración de N2O disuelto se determinó utilizando
un cromatógrafo de gases Bruker® GC-450 provisto de un
detector de captura electrónica, tomando unos 25 g (±0,01
g) de la muestra mediante el uso de una jeringa de cristal
(Agilent P/N 5190-1547) y 25 mL de un gas patrón de
concentración conocida (300 ppbv). Tras esto, se agita la
jeringa durante 5 minutos (VIBROMATIC Selecta) y se
deja reposar para alcanzar una situación de equilibrio. Por
último, el gas es inyectado en el cromatógrafo de gases.
Esta operación se realizó por duplicado para cada frasco
Winkler.
La concentración de gases en el agua se calculó a través de
las medidas realizadas sobre el espacio de cabeza de las
muestras, usando las solubilidades propuestas por Weiss y
Price (1980) [5].
Para la estimación de los flujos atmósfera-océano se utilizó
la expresión:
F = k(CW – C*)
-1
donde k (cm h ) es la velocidad de transferencia del gas,
Cw (mol L-1) es la concentración del gas en el agua, y C*
102
(2016)
(mol L-1) es la solubilidad del gas a la temperatura de
equilibración (25 ± 1 °C) y a la salinidad de la muestra.
RESULTADOS Y DISCUSIÓN
Durante los dos años, se ha observado una variabilidad
estacional de la concentración de N2O, además de la
misma distribución en todas las secciones, valores más
bajos en superficie y más elevados en zonas profundas. Sin
embargo, apenas se han apreciado variaciones
longitudinales de este gas. Esta distribución coincide a la
encontrada por Han et al (2013) [6] en el noroeste del Mar
del Sur de China. Tanto en aguas superficiales, como en
profundas, se ha observado un control térmico sobre la
concentración de N2O, existiendo una relación inversa
entre la distribución del N2O y la temperatura (Fig. 1). Este
mismo comportamiento, fue encontrado por Morell et al
(2001) [7], en un estudio realizado en las costas de Puerto
Rico.
Ferrón et al (2010a) [8], 7,3 Gg año-1, para una superficie
menor del Golfo de Cádiz (15,86 x 102 Km2). Los océanos
actúan como fuentes de N2O a la atmósfera, representando
la emisión oceánica de este gas, incluyendo las plataformas
continentales y los estuarios, el 29% de las emisiones
globales de N2O a la atmósfera [6].
Tabla 1. Variación de los flujos medios de N2O para cada sección:
Guadalquivir (GD), Sancti Petri (SP) y Trafalgar (TF),
diferenciando entre zona costera y distal.
Sección
Flujo N2O
(µmol m-2d-1)
Zona costera Zona distal
(< 75m)
(>75 m)
GD
3,82
3,69
SP
3,09
1,30
TF
2,21
0,62
AGRADECIMIENTOS
Este trabajo ha sido financiado por los proyectos STOCA
(Instituto Español de Oceanografía) y CTM2014-59244C3.
REFERENCIAS
Fig. 1. Variación de la concentración de N2O (nM) para la sección
de Guadalquivir durante junio de 2015.
No se han encontrado variaciones importantes de la
concentración de N2O entre las diferentes secciones
estudiadas, aunque la de Guadalquivir presenta valores
ligeramente superiores (10,12 ± 1,02 nM). Este hecho, se
puede deber a un mayor aporte continental de materia
orgánica y nutrientes proveniente del río Guadalquivir. Las
concentraciones globales medidas (9,98 ± 0,89 nM), son
más bajas que las determinadas por Ferrón et al (2010 a)
[8] (16,58 ± 0,89 nM) en aguas del Golfo de Cádiz, ya que
este último estudio se centró en una zona más próxima a
costa.
Los flujos medios de N2O son positivos, es decir, el Golfo
de Cádiz actúa como fuente de estos gases a la atmósfera,
exceptuando diciembre de 2015 que presenta un flujo
negativo, actuando así la zona como sumidero de dicho
gas. Los flujos de este gas, al igual que ocurre con sus
concentraciones, presentan una variabilidad estacional, con
valores más elevados en las estaciones de verano y otoño y
menores en invierno. A su vez, los flujos de N2O, presentan
valores mayores cerca de costa y menores en las zonas
distales en todas las secciones y para todo el periodo de
tiempo estudiado (Tabla 1). Este hecho podría atribuirse al
efecto que presentan los mayores aportes continentales y
fluviales en zonas más someras, produciendo así una
intensificación de los procesos de mineralización de la
materia orgánica con la correspondiente liberación de
nutrientes y gases a la columna de agua.
Las emisiones globales para el área de estudio (43,83 x 102
Km2) son de 0,21 y 0,17 Gg N2O año-1 en 2014 y 2015
respectivamente. Este valor es inferior al calculado por
1 - Houghton, J.T., Jenkins, G.J., Ephramus, J.J., 1990.
Climate Change: The Ipcc Scientific Assessment.
Cambridge University Press, Cambridge. 365 pp.
2 - Intergovernmental Panel of Climate Change (IPCC),
2014. Climate Change 2014: Synthesis Report. Summary
for Policymakers. 31 pp.
3 - Intergovernmental Panel of Climate Change (IPCC),
2013. Climate Change 2013: The Physical Science Basis.
Contribution of Working Group I to the Fifth Assessment
Report of the IPCC. [Stocker, T.F., D. Qin, G.-K. Plattner,
M. Tignor, S.K. Allen, J. Boschung, A. Nauels, Y. Xia, V.
Bex & P.M. Midgley (eds.)]. Cambridge University Press,
Cambridge, United Kingdom and New York, NY, USA,
1535 pp.
4 - Freing, A., Wallace, D. W., y Bange, H. W., 2012.
Global oceanic production of nitrous oxide. Philosophical
Transactions of the Royal Society B: Biological Sciences,
367(1593): 1245-1255.
5 - Weiss, R. F., y Price, B. A., 1980. Nitrous oxide
solubility in water and seawater. Marine Chemistry, 8(4):
347-359.
6 - Han, Y., Zang, J.Z., Zhao, Y.C., Liu, S.M., 2013.
Distributions and sea-to-air fluxes of nitrous oxide in the
coastal and shelf waters of the northwestern South China
Sea. Estuarine Coastal and Shelf Science, 133: 32-44.
7 - Morell, J. M., Capella, J., Mercado, A., Bauzá, J., y
Corredor, J. E., 2001. Nitrous oxide fluxes in Caribbean
and tropical Atlantic waters: evidence for near surface
production. Marine chemistry, 74(2): 131-143.
8 - Ferrón, S., Ortega, T., y Forja, J. M., 2010a. Nitrous
oxide distribution in the north-eastern shelf of the Gulf of
Cádiz (SW Iberian Peninsula). Marine Chemistry, 119(1):
22-32.
103
(2016)
9 - Seitzinger, S.P., Kroeze, C., Styles, R.V., 2000. Global
distribution of N2O emissions from aquatic systems:
natural emissions and anthropogenic effects. Chemosphere:
Global Change Sci., 2, 267-279.
104
(2016)
Historical record and sources of metals in core sediments from Maó
Harbour, Minorca, Spain
Antonio Tovar-Sánchez1,2, Marly C. Martínez-Soto3, David Sánchez-Quiles2, Jordi GarcíaOrellana4,5, Antoni Jordi3, Miguel A. Huerta-Diaz6, Gotzon Basterretxea3
1
Department of Ecology and Coastal Management. Andalusian Institute for Marine Science, ICMAN (CSIC). Campus
Universitario Río San Pedro, 11510 Puerto Real, Cádiz. Spain.
2
Department of Global Change Research. Mediterranean Institute for Advanced Studies, IMEDEA (CSIC-UIB), Miguel
Marqués 21, 07190 Esporles, Balearic Island, Spain.
3
Department of Ecology and Marine Resources, Mediterranean Institute for Advanced Studies IMEDEA, Universidad de las
Islas Baleares (UIB) - Consejo Superior de Investigaciones Científicas (CSIC), Esporles, Spain
4
Institut de Ciència i Tecnologia Ambientals, Universitat Autònoma de Barcelona, E- 08193 Bellaterra, Catalonia, Spain
5
Departament de Física, Universitat Autònoma de Barcelona, E-08193 Bellaterra, Catalonia, Spain
6
Instituto de Investigaciones Oceanologicas. Universidad Autonoma de Baja California. Campus Ensenada, Mexico
ABSTRACT
Maó (Minorca) is a narrow and semi-enclosed harbour impacted by historic urban and industrial metal pollution.
We analyse and compare surficial and long-term sediment records along the Harbour to assess the sources and
historical trends in pollution. Trace metal concentrations (Al, Ca, Co, Cr, Cu, Fe, Hg, Mn, Ni, Pb, Sr and Zn)
measured in sediments of Maó were of the same order of magnitude than those reported in other
anthropogenically impacted areas of the Mediterranean Sea. Mercury was the only element that presented
significantly higher concentrations in the harbour than other sites (up to 2 orders of magnitude higher). Although
surficial concentrations of some metals (e.g. Hg, Pb, Cu and Zn) showed levels lower than those presented in the
oldest sedimentary record, they showed concentrations above those recommended for acceptable sediment quality.
Enrichment factors (EF) calculations support the anthropogenic origin of these metals in the Maó harbour;
however, some variations can be attributed to natural fluctuations in the sediment deposition in the harbour.
INTRODUCTION
MATERIALS AND METHODS
Vertical profiles of metals in sediment cores have been
commonly used as "environmental records", providing
information on the current system and the geochemical
changes that occurred over time in the environment.
Chemical composition of sediment cores can be used to
establish background conditions and to evaluate how
contaminant levels have responded to changes in
population, land use and human activities [1-4].
Since the second half of the 19th century and especially
along the 20th century, the harbour’s sediments at Maó
became a reservoir of the waste generated along its shore,
where a continuous increase of demography,
industrialisation, construction, tourism and commerce took
place (Garcia-Orellana et al 2011). Here we reconstruct the
historical contamination of selected elements (Al, Cu, Co,
Cr, Fe, Hg, Mn, Ni, Pb, Zn, Ca and Sr) in Maó Harbour
using dated sediment cores as archives.
Core sediments (up to 55 cm depth) were collected by
scuba divers in June 2010 and July 2011 at 10 stations
located along the Maó Harbour (Figure 1). Radiometric
analyses (210Pb, 137Cs, 226Ra) were carried out on the bulk
fraction of sediments following the method described by
Garcia-Orellana et al. (2011) [4]. Metals were determined
by ICP-OES (Perkin Elmer ICP-OES Optima 5300 DV)
previous a microwave acid digestion according to the SW846 EPA Method 3051A [5], which involved the digestion
of 0.2 g of sediment sample by triplicate with 10 ml of
nitric acid (65%, Suprapur quality). Mercury
concentrations were determined by a Direct Mercury
Analyzer (Milestone DMA-80).
105
(2016)
(with high enrichment factors), most probably associated to
the resurgence of the jewellery industry and the
introduction of leaded gasoline in the early twentieth
century.
ACKNOWLODGEMENTS
This work was financed by the Ministerio de Economía y
Competitividad (MINECO) grants EHRE (CTM200908270). M.C. Martínez-Sotos’s work was funded by a JAEdoc contract from CSIC.
REFERENCES
Figure 1. Map showing the location of the study area in
Minorca Island (north-western Mediterranean Sea), and the
sampling stations corresponding to surface (S-) and cores
(C-) sediments.
RESULTS AND DISCUSSION
The length of the collected sediment cores, which ranged
from 22 cm (C-3) to 55 cm (C-5), provided information
from the 1890's up to the present. Metal concentrations
measured in surficial sediments of Maó Harbour were
similar to those reported in other anthropogenically
impacted Mediterranean Sea bays, except for Hg which
showed concentrations of up to 100 times higher. Our
results showed that the 1890s and 1900s were characterized
by high concentrations of Ca and Sr, indicating the
presence of carbonated algal communities in the harbour
waters, a characteristic that suggest absence of strong
anthropogenic influence. The 1910s, 1920s and 1930s
displayed marked increase of Hg and Pb concentrations
1. Dassenakis, M., M. Scoullos & Gaitis, A. 1997. Trace
metal transport and behaviour in the Mediterranean estuary
of Acheloos river. Mar. Poll. Bull. 34 (2): 103-111.
2. Rubio, B., Nombela M.A., & Vilas, F. 2000. Heavy
metal pollution in the Galician Rías Baixas: new
background values for Ría de Vigo (NW Spain) Journal of
Iberian Geology, 26, 121-149.
3. Tuncer, G., G. Tuncel & Turgut. B. 2001. Evolution of
metal pollution in the Golden Horn (Turkey) sediments
between 1912 and 1987. Mar. Poll. Bull. 42 (5): 350-360.
4. Garcia-Orellana, J. Cañas L., Masqué P., Obrador B.,
Olid C., Pretus J. 2011. Chronological reconstruction of
metal contamination in the Port of Maó (Minorca, Spain).
Mar. Poll. Bull. 62, 1632–1640.
5. US Environmental Protection Agency. 1987. An
overview of sediment quality in the United States. EPA
905/9-88-002. Office of Water Regulations and Standards,
Washington, DC, and EPA Region 5, Chicago
106
(2016)
Distribution and transport of dissolved trace metals in the Gulf of Cádiz,
Spain
Antonio Tovar-Sánchez1,2, David Sánchez-Quiles2, David Roque1, Antonio Cobelo, Irene
Laiz3, Ricardo Sánchez, Miguel Bruno
1
Department of Ecology and Coastal Management. Andalusian Institute for Marine Science, ICMAN (CSIC). Campus
Universitario Río San Pedro, 11510 Puerto Real, Cádiz. Spain.
2
Department of Global Change Research. Mediterranean Institute for Advanced Studies, IMEDEA (CSIC-UIB), Miguel
Marqués 21, 07190 Esporles, Balearic Island, Spain.
3
Department of Applied Physics, University of Cadiz, Campus Rio San Pedro, 11510, Puerto Real, Cádiz, Spain
ABSTRACT
The Gulf of Cadiz plays a key role in the exchange of biogeochemical fluxes between the Mediterranean Sea and
the Atlantic Ocean through the Strait of Gibraltar. Oceanographers have carried out many investigations in the
Gulf of Cádiz on water mass circulation and mass balance of nutrients and carbon. However, despite its
importance in the global ocean functioning, studies on trace metals in the Gulf of Cádiz and Strait of Gibraltar
waters are very scarce. Here we show the concentrations of dissolved trace metal composition (i.e. Ag, Cd, Co,
Cu, Fe, Mo, Ni, Pb and Zn) in the Gulf of Cádiz and Mediterranean Sea surface waters as obtained from 5
oceanographic campaigns. Our results indicate that the Gulf of Cádiz surface water mass is receiving large
amounts of trace metals transported by the different rivers that flow into the Gulf of Cádiz. Thus, dissolved trace
metals in these waters were highly variable with the highest ranges measured for Co (0.06 – 3.1 nM), Fe (0.6 –
392 nM) and Pb (0.04 – 512 nM).
INTRODUCTION
The Gulf of Cádiz is a semi-enclosed basin with an
oceanographic dynamics mainly controlled by the
exchanges between the Mediterranean Sea, the Atlantic
Ocean, the coastal system, the atmosphere and the seafloor.
The Gulf of Cádiz water masses are directly influenced by
the Iberian pyrite belt (one of the largest sulfide deposits in
the world), receiving large amounts of trace metals
transported by the different rivers that run through the belt
and flow into the Gulf of Cádiz [1-4]. Trace metals play a
critical role in the ocean functioning. Some metals, e.g. V,
Cr, Mn, Fe, Co, Ni, Cu, Zn or Mo, despite of being present
in organisms at trace concentrations, are considered
essential for life if ambient concentrations do not exceed a
threshold value for toxicity [5]. On the other hand, the
presence of some trace metals at high concentrations such
as As, Hg, or Pb can damage the ecosystem health. Thus,
the distribution of trace metals can enhance or limit
primary productivity in some regions of the world ocean.
Despite their importance, we do not have a good
understanding of the global distribution and cycling of
trace metals in many regions of the ocean.
We collected surface water samples during 5
oceanographic cruises carried out during 2014 – 2015 in
the Gulf of Cadiz and the Mediterranean Sea, providing an
opportunity to further advance in the knowledge of both
regions, as well as of the distribution of trace elements
within the Gulf of Cádiz surface waters and their inflow to
the Mediterranean Sea.
MATERIALS AND METHODS
The sampling was carried out on board different
oceanographic vessels, during October and December
2014, and during March, September and November 2015
(Figure 1). Samples for trace metals were collected using a
teflon tow-fish sampling system deployed at approximately
2 m depth utilizing established trace metal-clean
techniques. After sample collection, the seawater was
filtered on board through acid-washed 0.2 µm filter
cartridges and acidified using Optima grade HCl to a
pH<2. Dissolved samples (<0.2 μm) were double bagged in
polyethylene bags and shipped to the trace metal clean
laboratories, where they were preconcentrated by an
organic extraction procedure using the APDC/DDDC
ligand technique [6]. The levels of metals were quantified
by ICP-MS. To evaluate the accuracy of our analytical
procedures, a certified seawater reference material (CASS5) was preconcentrated and analyzed with the samples.
RESULTS AND DISCUSSION
107
(2016)
Dissolved metal (Ag, Cd, Co, Cu, Fe, Mo, Ni, Pb and Zn)
concentrations measured in surface waters during our study
varied broadly with the geographic location. The highest
concentrations of most of the metals were measured in
areas under the influence of the rivers discharge (e.g. Ag:
11.3 pM, Cd: 0.3 nM, Co: 2.1 nM, Cu: 13.3 nM, Fe: 88.6
nM, Mo: 121.1 nM, Ni: 70.8 nM and Zn: 49.5 nM within
the Guadalquivir river area of influence during December
2014). On the other hand the lowest concentrations were
measured in offshore waters (e.g. Ag: <1 pM, Cd: 0.17 nM,
Co: 60 pM, Cu: 1.9 nM, Fe: 0.6 nM, Mo: 105.5 nM, Ni: 2.3
nM and Zn: 1.1 nM).
Currents data were collected with a vessel-mounted ADCP
along the ship track (see Fig. 1). At each sampling
location, the mean surface alongshore and cross-shore
velocities were obtained by spatially averaging the first top
valid bin between two consecutive sampling locations.
Thus, the horizontal advective fluxes of metals can be
estimated by multiplying each velocity component by the
different metal concentrations.
Figure 1. Sampling locations for the different
oceanographic cruises
ACKNOWLODGEMENTS
This work was financed by the MICCIN grants MEGOCA
(CTM2014-59244-C3-3-R). Thanks also to Joaquin
Pampin and Antonio Moreno for their collaboration during
sampling preparation.
REFERENCES
[1]. Van Geen A., Boyle E.A., Moore W.S. 1991. Trace
metal enrichments in waters of the Gulf of Cadiz, Spain.
Geochimica et Cosmochimica Acta 55, 8, 2173-2191
[2]. Elbaz-Poulichet F., Morley N.H., Cruzado A.,
Velasquez Z., Achterberg E.P., Braungardt C.B. 1999.
Trace metal and nutrient distribution in an extremely low
pH (2.5) river-estuarine system, the Ria of Huelva (SouthWest Spain). The Science of the Total Environment 227,
73-83
[3]. Elbaz-Poulichet, F., Morley, N., Beckers, J.M.,
Nomerange, M., 2001a. Metal fluxes through the Strait of
Gibraltar: the influence of the Tinto and Odiel rivers (SW
Spain). Mar. Chem. 73, 193–213.
[4]. Elbaz-Poulichet F., Braungardt C., Achterberg E.,
Morley N., Cossa D., Beckers J-M, Nomérange P.,
Cruzado A., Leblanc M. 2001b. Metal biogeochemistry in
the Tinto–Odiel rivers (Southern Spain) and in the Gulf of
Cadiz: a synthesis of the results of TOROS Project.
Continental Shelf Research 21, 1961–1973
[5]. F. J. Stevenson y M. A. Cole, Cycles of soil. Carbon,
Nitrogen, Phosphorus, Sulfur, Micronutrients, Second
edition. Wiley, 1999
[6]. Tovar-Sánchez, A. 2012.“1.17 sampling approaches
for trace element determination in seawater” in
Comprehensive Sampling and Sample Preparation, ed
J.Pawliszyn (Oxford:AcademicPress), 317–334.
108
(2016)
Contaminación de sistemas costeros por edulcorantes artificiales: fuentes,
distribución y persistencia
Juan M. Traverso-Soto1, Rosa M. Baena-Nogueras1, Miriam Biel-Maeso1 & Pablo A. LaraMartín1
1
Departamento de Química Física, Facultad de Ciencias del Mar y Ambientales, CEI-MAR, Universidad de Cádiz, Campus
Río San Pedro, Puerto Real, Cádiz, 11510.
RESUMEN
Durante la última década se ha incrementado el uso de edulcorantes artificiales en la dieta como sustituto del
azúcar. Una vez consumidos, son excretados sin metabolizar y su eliminación en estaciones depuradoras de aguas
residuales (EDARs) es también deficiente. Esto ha llevado a su detección en el medio ambiente, donde aún se
desconocen sus efectos. En este sentido hemos realizado una serie de muestreos en la Bahía de Cádiz para medir
la concentraciones de los 6 edulcorantes con uso autorizado en la UE: sucralosa (SUC), sacarina (SAC),
aspartamo (ASP), acesulfamo (ACE), ciclamato (CYC) y neohesperidina dihidrocalcona (NHDC). Las máximas
concentraciones detectadas fueron para SUC (> 2000 ng/L), seguido de ACE (290 ng/L), en el estuario del
Guadalete, ambos compuestos presentando un comportamiento conservativo a lo largo de esta área. Su
persistencia, confirmada en ensayos de degradación en el laboratorio, los convierte en candidatos para su uso
como trazadores de contaminación por aguas residuales en el medio marino.
INTRODUCCIÓN
Un edulcorante artificial es una sustancia sintética
empleada como sustituto del azúcar fundamentalmente por
su mayor dulzura y menor aporte calórico. Su producción y
uso está en alza (ej.: incrementos superiores al 5% en los
últimos años en EE.UU.), particularmente asociado al
consumo de productos dietéticos [1]. Hasta la fecha no se
ha prestado mucha atención a la presencia y
comportamiento de estos y otros ingredientes alimentarios
en el medio ambiente. En el caso de los edulcorantes
artificiales se trata de compuestos aniónicos con elevada
solubilidad acuosa que han sido recientemente detectados
en aguas residuales (tanto influentes como efluentes de
EDARs, donde presentan bajos porcentajes de
eliminación), aguas subterráneas y ríos [2]. Más
recientemente, se han medido concentraciones de hasta 32
ng/L en el Mar del Norte [3] para el caso de la sucralosa
(SUC).
El objetivo de este estudio es incrementar el conocimiento
sobre el comportamiento de estos aditivos en el medio
marino. Concretamente, se trata de analizar la presencia de
los edulcorantes artificiales más usados en la UE (SUC,
SAC, ASP, ACE, CYC y NHDC) en aguas costeras,
seleccionando para ello diversos puntos de muestreo de
aguas superficiales en la Bahía de Cádiz (SO de España).
Además de conocer su distribución espacial, se pretende
establecer si se trata de sustancias persistentes en el medio
marino mediante la realización de ensayos de foto y
biodegradación en el laboratorio.
MATERIAL Y MÉTODOS
Durante la pleamar se recogieron muestras superficiales de
agua de mar en distintos puntos de la Bahía de Cádiz (Fig.
1) mediante botellas de vidrio ámbar de 2.5 L. Tras su
filtración con filtros de fibra de vidrio de 1 micra de
tamaño de poro, se procedió al aislamiento de los
edulcorantes mediante extracción en fase sólida (SPE)
usando cartuchos Oasis HLB 500 mg y siguiendo una
modificación del protocolo desarrollado por Ordoñez et al.
[4]. La determinación de los analitos se hizo mediante
cromatografía líquida de ultra resolución acoplada a
espectrometría de masas en tándem (UPLC-MS/MS). La
metodología empleada arrojó recuperaciones superiores al
75% para los compuestos analizados y límites de detección
inferiores a 1 ng/L.
Adicionalmente, se realizaron ensayos de biodegradación y
fotodegradación siguiendo las guías OECD 306 y 316,
respectivamente. En el primer caso se incubó agua de mar
del área de muestreo a 18ºC durante 28 días en oscuridad,
adicionándose edulcorantes al inicio (1 ng/mL) y
analizándose muestras a intervalos regulares. Este agua,
tras esterilización mediante filtración por 0.22 micras, se
usó para experimentos de fotodegradación, con una
duración de 24 h. En este caso se usó un fotorreactor
Suntest CPS+ irradiando a 500 W/m2 y tubos de cuarzo
donde la disolución de agua de mar y edulcorantes (10
ng/mL) se analizó a intervalos regulares.
109
(2016)
La persistencia de estos dos edulcorantes quedó también
puesta de manifiesto al estudiar el perfil longitudinal de
concentración de los mismos a lo largo del río Guadalete
(Fig. 2), el cual muestra claramente un comportamiento
conservativo para ACE y SUC. En este sentido, se
corrobora su idoneidad como trazadores de contaminación
por aguas residuales de tipo urbano en sistemas acuáticos,
previamente observada en aguas subterráneas y ríos [2].
Fig. 1. Mapa de la Bahía de Cádiz mostrando la
localización de las distintas estaciones de muestreo en 3
áreas (Estuario Guadalete, Río San Pedro, Caño Sancti
Petri).
RESULTADOS Y DISCUSIÓN
En la Tabla 1 se muestran los intervalos de concentraciones
medidos en aguas superficiales de la Bahía de Cádiz. La
zona que presenta una mayor contaminación es del estuario
del río Guadalete, alcanzándose concentraciones máximas
(> 2000 ng/L) aguas arriba, en las proximidades de la
EDAR Jerez del Frontera, uno de los más notables focos de
contaminación en la bahía [5]. Se puede observar como la
sucralosa (SUC) es el compuesto que presenta una mayor
concentración, seguido de acesulfamo (ACE). La mayor
concentración en el medio ambiente de estos compuestos
es consecuencia tanto de su mayor volumen de uso frente a
otros edulcorantes (ej.: NHDC, no detectado en ninguna de
las muestras) como debido a su carácter persistente, ya que
durante los ensayos de biodegradación (28 días) y
fotodegradación (24 h) realizados con agua del área de
muestreo no se observó disminución significativa en su
concentración.
Tabla 1. Concentración de edulcorantes (ng/L) en aguas
superficiales de la Bahía de Cádiz.
Estuario
Río San
Caño
Guadalete
Pedro
Sancti Petri
SUC
512-2174
517-1536
520-1199
SAC
13-83
5-22
n.d.-22
ASP
8-37
7-26
7-26
ACE
17-290
9-24
n.d.-6
CYC
17-51
10-20
2-14
NHDC
n.d.
n.d.
n.d.
Fig. 2. Concentración de edulcorantes (ACE = acesulfamo,
SUC = sucralosa) versus salinidad en el estuario del Río
Guadalete.
AGRADECIMIENTOS
La financiación de esta investigación ha sido realizada
mediante el proyecto RNM 6613 de la Consejería de
Innovación, Ciencia y Empresa de la Junta de Andalucía.
REFERENCIAS
1 – Sylvetsky, A.C., Rother, K.I., 2015. Trends in the
consumption of low-calorie sweeteners. Physiology &
Behavior (in press).
2 – Buerge, I.J., Buser, H.R., Kahle, M., Müller, M.,
Poiger, T., 2009. Ubiquitous occurrence of the artificial
sweetener acesulfame in the aquatic environment: an ideal
chemical marker of domestic wastewater in groundwater.
Environ. Sci. Technol. 43:4381-4385.
3 – Brumovsky, M., Becanova, J., Kohoutek, J., Thomas,
H., Petersen, W., Sorensen, K., Sanka, O., Nizzetto, L.
2016. Exploring the occurrence and distribution of
contaminants of emerging concern through unmanned
sampling from ships of opportunity in the North Sea. J.
Mar. Sys. (in press).
4 - Ordoñez, E.Y., Quintana, J.B., Rodil, R., Cela, F., 2013.
Determination of artificial sweeteners in water samples by
solid-phase extraction and liquid chromatography–tandem
mass spectrometry. J. Chromatogr. A., 1256:197-205.
5 – Lara-Martín, P.A., Gómez-Parra, A., González-Mazo,
E., 2008. Sources, transport and reactivity of anionic and
non-ionic surfactants in several aquatic ecosystems in SW
Spain: a comparative study. Environ. Poll. 156:36-45.
110
(2016)
Oxidative stress and neurotoxicity in Scrobicularia plana exposed to
pharmaceutical mixture
Chiara Trombini1, Miriam Hampel2 & Julián Blasco1
1
Instituto de Ciencias Marina de Andalucía (CSIC), Campus Río San Pedro, 11510 Puerto Real, Cádiz, España 2 Centro
Andaluz de Ciencias y Tecnologías Marinas (CACYTMAR), Universidad de Cádiz, Campus Universitario de Puerto Real,
11510 Puerto Real, Cádiz, España
ABSTRACT
Pharmaceuticals are pollutants of potential concern in the aquatic environment where they are commonly
introduced as a complex mixture. In this study we have evaluated the toxic effects produced in clams
Scobicularia plana by a mixture of ciprofloxacine, flumequine and ibuprofen, three pharmaceuticals that are
produced and used in large quantity and have a wide spread occurrence in aquatic environment. Clams were
exposed to two mixture concentrations (at 10 and 100 µg·L-1 levels) during 28 days (21 days of exposure and 7
of depuration) and oxidative and neurotoxic effects were assessed using a multi-biomarker approach: CAT, GR,
GPx, GST, SOD, LPO levels and AchE activity were evaluated in gills and digestive gland over time. The
present study demonstrated that the combination of three pharmaceutical compounds, particularly at the highest
concentration tested, have a considerable effect on the activities of antioxidant and detoxifying enzymes and
therefore on the oxidative status of S. plana.
INTRODUCTIÓN
Nowadays, the presence of pharmaceutical compounds in
marine compartments is well documented [1,2] and studies
to know effects of these emerging contaminants on aquatic
organisms have increased significantly in the last decades
[3-5]. However, due to the huge variety of pharmaceutical
compounds produced worldwide, toxic effects of most of
them are almost unknown.
Ciprofloxacin (CIP) and flumequine (FL) are broadspectrum antibiotics of the fluoroquinolones class which
exert their bactericidal effects by inhibiting the bacterial
DNA gyrase and therefore DNA replication.
Fluoroquinolones toxicity was observed in mice where
oxidative stress, cyto and neurotoxicity were induced after
treatment with CIP (0.5-300 mg·L-1 during 24, 48, 72 y 96
h; 10 µM-1.0 mM during 72 h) [6,7] and
hepatocarcinogenesis probably related to oxidative stress
was observed in mice treated with 4000 ppm FL during 6
weeks [8,9]. Both have been detected in different aquatic
compartments (e.g. rivers, effluents of WWTPs, hospital
effluents, groundwater) in concentrations ranging between
ng·L-1 and µg· L-1 [10-12].
Ibuprofen (IBU) is one of the most used non-steroidal antiinflammatory drugs (NSAIDs): its effects have been
studied in different marine organisms and its toxicity at
environmentally relevant concentrations (oxidative stress
induction, endocrine disruption, cyto and genotoxicity) has
been widely proven [13-15].
The aim of this study was to assess the chronic toxicity of a
mixture of CIP, Fl and IBU on the clam Scrobicularia
plana. Organisms were exposed to two mixture
concentrations of pharmaceuticals (10 and 100 µg·L-1)
during 28 days and, to evaluate toxic effects of mixture, a
multi-biomarker approach including antioxidant enzyme
activities (CAT, GR, GPx, GST, SOD), LPO levels and
AchE activity was chosen
MATERIALS AND METHODS
Specimens of S. plana collected in a reference site (San
Pedro River, Cádiz, Spain) were acclimated during 1 week
before starting the experiment. Exposure was carried out in
tanks with 0.45 µm filtered seawater (1 L/animal), aeration
and in semi-static conditions (complete water and chemical
renovation every 48 h). Pharmaceuticals were added as a
stock solution prepared in DMSO to reach final
concentration of 10 and 100 µg·L-1 (for each compound
present in the mixture). Clams were exposed for 28 days
(21 of exposure and 7 of depuration) and sampling times
were set up: 0 (control 0), 1, 7, 21 and 28 days. Triplicates
were used for each condition (seawater control, DMSO, 10
µg·L-1 and 100 µg·L-1 mixture). Temperature (18.0 ± 1.5
°C), salinity (31.0 ± 0.3 ppt), pH (7.9 ± 0.4) and dissolved
oxygen (8.8 ± 0.6 mg·L-1) were checked daily.
Organisms sampled over each selected time were dissected
and tissues (gills and digestive gland) quickly frozen in
liquid nitrogen and stored at -80 °C until further analysis.
Tissues were adequately treated (homogenization in 50
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mM Tris-HCl with 150 mM NaCl, 1 mM DTT, 0.1%
antiproteolitic cocktail and centrifugation at 12000 x g, 4
°C during 30 min) to obtain S12 fraction that was used to
quantify total proteins, antioxidant enzyme activity (CAT,
GR, GPx, GST, SOD), LPO and AchE activity. All
analyses were performed by spectrophotometry.
DISCUSSION
To date, few studies have been carried out to investigate
CIP and FL toxicity on aquatic organisms [16]. However,
toxic effects of IBU are studied in different marine species
and its ability to induce general stress and damage to
different levels in organism is well shown [15, 17]. No
studies have examined the joint effects of these
compounds. Activity of enzymes related to oxidative stress
was enhanced after exposure to the pharmaceutical
mixture, particularly at the highest concentration. Higher
activities were reported in digestive gland than in gills
indicating a greater importance of this organ in
detoxification processes.
AKNOWLEDGEMENTS
This work has been funded by the Ministerio de Economia
y Competitividad of the Spanish Government under Project
CTM2012-38720-C03-03, and ERF funds.
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