Studying the Effect of CO2-Induced ... - ACS Publications

Article pubs.acs.org/est

Studying the Effect of CO2‑Induced Acidification on Sediment Toxicity Using Acute Amphipod Toxicity Test M. Dolores Basallote,*,† Manoela R. De Orte,†,‡ T. Á ngel DelValls,† and Inmaculada Riba† †

Cátedra UNESCO/UNITWIN WiCop. Departamento de Química-Física, Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz, Polígono Río San Pedro s/n, Puerto Real, Cádiz 11510, Spain S Supporting Information *

ABSTRACT: Carbon capture and storage is increasingly being considered one of the most efficient approaches to mitigate the increase of CO2 in the atmosphere associated with anthropogenic emissions. However, the environmental effects of potential CO2 leaks remain largely unknown. The amphipod Ampelisca brevicornis was exposed to environmental sediments collected in different areas of the Gulf of Cádiz and subjected to several pH treatments to study the effects of CO2-induced acidification on sediment toxicity. After 10 days of exposure, the results obtained indicated that high lethal effects were associated with the lowest pH treatments, except for the Riá of Huelva sediment test. The mobility of metals from sediment to the overlying seawater was correlated to a pH decrease. The data obtained revealed that CO2-related acidification would lead to lethal effects on amphipods as well as the mobility of metals, which could increase sediment toxicity.

■

from subseabed storage sites.8 In addition, CO2 flows could reach distances of more than tens of kilometers from the storage site, depending on the thermodynamic state of the substance as well as the amount, velocity, and duration of release and even depend on the weather conditions.9 Thus, coastal and estuarine areas are likely susceptible to be impacted by acidification processes. In this sense, there remains a lack of experience with safety issues surrounding this technology which leads to uncertainties in risk assessments regarding the CCS operational procedure.10 Although catastrophic releases of CO2 are unlikely, the complexity of the sequestration and storage processes include a high degree of uncertainty that could bring the environment under unprecedented changes.11 One of the main effects expected if CO2 escapes through the sediment would be the acidification of the surrounding environment and pH reductions. In addition to the direct effects of a decrease of pH and increase in dissolved carbon dioxide (pCO2), changes related to CO2 seepages would most likely lead to changes in the biogeochemistry in the sediment−water interface, such as, metal mobilization, the leaching of nutrients and modification of protons gradients through biological membranes.12,11,13 A CO2 leakage may cause rapid extraction of easily soluble fractions of some trace elements or slow enhancement of easily soluble fractions of other trace elements.

INTRODUCTION Carbon capture and storage (CCS) is considered one of the better choices for the near-term reduction of atmospheric CO2 emissions required for the UNFCCC by 2025 and 2050 (15% and 60% of actual emissions, respectively).1 Conventions protecting the oceans against pollution establish that CO2 emitted by power plants could be stored deep below the sea bottom.2,3 This mitigation strategy is based on the liquefaction, transport, and storage of CO2 as a supercritical state into subseafloor porous rocks. Deep saline aquifers, depleted oil and gas reservoirs, and unmineable coal appear to be the most common choices for storing the CO2.4 Because of the low probability of a large-scale CO2 leakage, this technology is being considered economically, socially, and environmentally feasible by expert opinions in the fight against climate change.5 This technology has increased 50% since 2011. About 40 million tons of CO2 are being captured per annum globally, and about 10 million are expected to be captured in the next few years by new and advanced-stage projects. Northern Europe and America are leading this technology.6 Furthermore, among all the CCS site selections, about half of the projects have been located at offshore areas. Specifically, in Spain, 11 areas around the national territory have been proposed as potential underground storage sites. Some of these proposed formations are located in coastal areas, where some zones will be likely contaminated by metals.7 Transport pipeline failure, faulty injection wells, faults or fractures in caprock, or seepages through porous geological structures are some of the possible mechanisms for CO2 escaping © 2014 American Chemical Society

Received: Revised: Accepted: Published: 8864

March 28, 2014 June 18, 2014 July 2, 2014 July 2, 2014 dx.doi.org/10.1021/es5015373 | Environ. Sci. Technol. 2014, 48, 8864−8872

Environmental Science & Technology

Article

mining activities within their basins.26−28 While the ML site is located in the inner part of the estuary in the Odiel River, the MAZ site is located at the mouth of the estuary. The water used in the experiments (SWRSP) was collected from the Rió San Pedro surface (1 m depth) during high tide, transported to the laboratory, and placed in a 400-L tank. The seawater was continuously filtered using a high-power external filter (CrystalProfi - e900) containing nitrate removal stones (Denitrate, Seachem). Sediment−Water Characterization. The surface sediments (0−5 cm layer) with their initial water content were collected directly using plastic 25-L containers at the Bay of Cádiz sampling sites (RSP, SF, TRO), because there was no water overlying the sediment at low tide, and by professional divers at the Riá de Huelva sampling sites (MAZ, ML) due to the seawater depth. The buckets were closed hermetically, taken to the laboratory, and stored in a cooler (4 °C) until the toxicity tests were performed, never longer than 2 weeks. The sediments were sieved through a 1 mm plastic mesh and homogenized with a Teflon spoon until no color or textural differences could be observed. All the beakers were thoroughly cleaned with acid (10% HNO3) and rinsed in double deionized (Milli-Q) water before the sampling and storage process. Sediment subsamples were collected for chemical quantification of organic carbon, organic matter content, grain size distribution, and metal concentration. The organic matter content (OM) in the sediment was determined through calcinations (LOI) in a muffle furnace at 450 °C. The total organic carbon content (TOC) was determined using the technique described by Gaudette et al.29 and modified by ElRayis.30 The grain size distribution was determined according to ASTMD422-6331 and Gee and Or.32 The percentage of sand and fines (silt and clay) was used for sediment textural classification, as described by Flemming.33 The metal concentrations (Fe, Cr, Cu, Ni, Co, Zn, Pb, Cd, and As) in the sediment were determined in the total fraction according to Loring and Rantala.34 The sediment samples (0.1 g of 40 °C dried sediment sample) were digested with a mixture of acids (nitric, chlorhydric, and fluorhydric acid) under controlled temperature (first at 180 °C, and posterior heating at 110 °C), and complexation with boric acid (10 min). After digestion by microwaves (Speedwave of Berghof), samples were analyzed using inductively coupled plasma−mass spectrometer (ICP-MS) (Thermo Elemental Series-X). Water samples were acquired from each test vessel at the end of the exposure time, filtered (0.45 μm) and acidified to pH < 2 with ultrapure grade HNO3 for posterior metals analysis. The concentrations of Fe, Cr, Cu, Ni, Co, Zn, Pb, Cd, and As were determined by inductively coupled plasma-mass spectrometry (ICP-MS) (Thermo Elemental Series-X). Metal analysis was performed in the Spectroscopy Division (ICP/AAS) ́ of the of the “Servicios Centrales de Ciencia y Tecnologia” University of Cádiz by an internal validated method.35 Polyatomic salt water interferences were controlled by coalition-reaction cell. Overlying seawater (50 mL) was collected and preserved in darkness, avoiding any head space for posterior analysis of total alkalinity (TA). TA was determined by automatic titration (Mettler Toledo, T50) using a combined glass electrode (Mettler Toledo, DGi115-SC) calibrated on the NBS scale. The TA and CO2 system pH were used to determine the seawater−carbonate system speciation. Data of total inorganic carbon (TIC), bicarbonate (HCO3−), carbonate (CO32−), the partial pressure of CO2 (pCO2), and saturation states for

The enhancement of soluble fractions of heavy metals may create toxic cascade effects in marine ecology, from benthic to pelagic systems.14 However, much work remains to be performed to understand the complex interrelationship between the various marine chemical reactions.15 Therefore, a better understanding of the mechanism of environmental responses to acidification processes at both deep-sea and shallow waters levels is needed when the effects of acidification processes on aquatic environments are being evaluated.12,16 Amphipods are marine organisms that are ecologically relevant. Because these organisms are widely distributed in aquatic environments, they play an important role in the food chain, as the principal food for fish and birds.17 These marine benthic organisms live in direct contact with sediments, and consequently, they are sensitive indicators of sediment pollution. These organisms are suitable for laboratory experiments because they have a short life cycle and are easily collected and manipulated. Standard protocols have described the use of amphipod crustaceans for whole sediment toxicity assessments.18−20 The sediment toxicity assay using the autochthon species in the South of Europe, A. brevicornis, has been interlaboratory calibrated for use in sediment quality assessment studies on the Atlantic Coast.21,22 The research work presented here attempts to evaluate the adverse effects of pH decreases associated with CO2 leaks from submarine geological structures on marine organisms. To increase the knowledge about the marine environmental risk related to CCS technologies, a nonpressurized CO2 injection system has been used to perform ecotoxicological assays. The CO2 injection system was used to study the effects of elevated CO2 levels on amphipod organisms by performing the 10-day standardized acute toxicity test. Furthermore, secondary effects, such as metal mobilization from the sediment to the overlying seawater, as a consequence of CO2 acidification events, were also studied to estimate other indirect effects linked to CO2 leakages.

■

MATERIALS AND METHODS Sediment Sampling Sites. Sediment samples were collected from different areas located along the Southern Atlantic Coast of Spain (see Supporting Information (SI) Figure S1). Rió San Pedro (RSP), Trocadero (TRO), and San Fernando (SF) are located in the Bay of Cádiz, while Mazagón (MAZ) and Muelle de Levante (ML) are located in the Riá of Huelva. The Bay of Cádiz is a shallow (less than 3 m) semienclosed system mainly affected by marine aquaculture, industrialization, urban discharges, and other anthropogenic activities such as the shipbuilding industry and car and aircraft manufacturing.23 The RSP sample site is a shallow tidal creek area exposed to west wind and subjected to a semidiurnal tidal regime, from 3.5 m at spring tide to 0.5 m at neap tide.24,25 The RSP sediment was considered a relatively uncontaminated reference site. The TRO and SF sites are located in the inner part of the Bay, which is a tidally controlled relatively protected area connected to the Atlantic Ocean through intertidal channels and salt-marshes. The main sources of contamination in this area come from the San Fernando city wastewater discharge, and it has been affected by industrial shipyards for years.23 The Riá of Huelva is a complex system of drainage streams that are tidally controlled, which receives the inputs of the Odiel and Tinto rivers and exchanges water directly into the open sea. This area is affected by chronic contamination from industrial sewage, urban sewage, and most importantly by fluvial inputs from the Odiel and Tinto rivers. These rivers are highly contaminated by metals due to long-term 8865

dx.doi.org/10.1021/es5015373 | Environ. Sci. Technol. 2014, 48, 8864−8872


Article

software, to determine significant differences (p < 0.01, p < 0.05) between the different pH treatments and sediments. Thus, for each test sediment (SF, TRO, and RSP), significant differences between the pHs of 7.5, 7.0, 6.5, and 6.0 and the sediment at its natural pH (nominal pH 8.0) were studied. In addition, for each pH treatment, the mortality differences between the sediments SF and TRO in reference to the RSP sediment test (considered the reference site) were calculated. The same analysis of variance (ANOVA) was employed to compare the metal mobility to the overlaying test water for the difference pH treatments with respect to the RSP sediment. The correlation between metal concentration in overlying test water and pH was determined. Normality and homogeneity of variance were tested, and the data were logarithmically transformed when necessary to satisfy the assumptions of ANOVA. The mortality was calculated based on the survival percentage measured after 10 days of exposure. These data were used to calculate the toxic parameter LpH50, which was defined as the pH that causes lethal effects in 50% of the population exposed. This parameter was estimated using the pH that provokes mortality in 50% of the population exposed to the different pH treatment and was calculated by the linear interpolation method for lethal toxicity. A factor analysis was conducted in order to explain the relationship observed between an initial number of variables, using a lower number of factors, by principal component analysis (PCA) at the software SPSS 15.0.

aragonite and calcite were calculated using the CO2SYS program36 with the dissociation constant from Mehrbach et al.37 with a refit by Dickson and Millero38 and KSO4 using Dickson.39 The pH in each aquarium was verified using a portable pH meter (model: Phenomenal 1000 H; accuracy ±0.005 pH), calibrated using pH buffer solutions of 4.00 and 7.00. The oxygen saturation was recorded using an Oxy 4000 H meter (accuracy ±0.5% of the measured value), and salinity was measured using a conductivity EC300 m (VWR, accuracy ±1% of the measured value). Acute Amphipod Toxicity Test. Acute amphipod toxicity tests using the specie A. brevicornis have been previously validated for studies developed on the Atlantic Coast.21,40−43 Several independent 10-day acute toxicity tests were developed under laboratory conditions using a nonpressurized laboratory-scale CO2 injection system. The system has been described in detail by Basallote et al.44 and De Orte et al.45 Briefly, this experimental device is composed of 12 test vessels, in which the pH is controlled independently by an automatic computer system (Aqua Medic AT Control) using CO2 injection. The pH in each experiment vessel was monitored continuously and adjusted by release of pure CO2 gas through a solenoid valve, which opens delivering the gas when an increase of pH of 0.01 units above the set value is detected. After the required pH value is reached, the solenoid valve is closed, stopping the input of CO2 in the test vessel. The CO2 injection system has been adapted to work with amphipod organisms using 2-L glass beakers. The amphipod organisms were collected from a low contaminated area at the Bay of Cádiz by sieving through a 0.5 mm mesh. Once in the laboratory, the organisms were acclimated to the laboratory conditions for 2 weeks before toxicity testing. Sediment toxicity tests were performed using 2-L glass beakers in a 1:4 v/v sediment water relation. The test vessels were left to settle and then left under constant aeration for 24 h before the amphipods were added. The test parameters and conditions, as well as the different pH treatments are summarized in SI Table S1. Aeration was provided with a Pasteur pipet without disturbing the sediment surface. The temperature, salinity, and dissolved oxygen were measured and controlled daily in each replicate. Randomly selected organisms (n = 20, per duplicate) were exposed to several sediment-seawater pH treatments. The pH treatments were selected according to (1) some predicted pH reductions scenarios caused by CO2 leakages in marine ecosystems and (2) the expected pH-related mortality for this species. Thus, the natural pH of coastal-estuarine area, which was used as the control sediment treatment where no CO2 was added, was compared with various simulated pH reductions due to leakages from CO2 storage sites. A set value of 0.5 U pH reductions were selected ranged between 8.0 and 6.0. Then, the pH treatments were identified as treatment 1 (pH 7.6 ± 0.1), treatment 2 (pH 7.1 ± 0.02), treatment 3 (pH 6.5 ± 0.04), and treatment 4 (pH 6.1 ± 0.03). The nominal pH treatments were 8.0 (control), 7.5, 7.0, 6.5, and 6.0, in that order. The pH was decreased slowly (by 0.5 U/day) by bubbling CO2 gas through a ceramic bubble diffuser placed at the base of each test vessel to avoid a sudden change of pH in the experiment vessels. Adequate quality assurance/quality control (QA/QC) measurements were followed in all aspects of the study, from field sampling to the laboratory as per DelValls et al.46 Statistical Analysis. The variation in mortality rates between pH treatments was examined using analysis of variance (ANOVA), followed by Dunnettś test, using SPSS 15.0 statistical

■

RESULTS AND DISCUSSION Physicochemical Parameters. The sediment and seawater physical and chemical characterization data are presented in SI Table S2. Regarding the fines percentages, the sediment from RSP, SF, and MAZ was classified as muddy sand, while that from SF and ML were classified as sandy mud. The highest percentages of TOC and OM were observed at ML and TRO, respectively. All the test sediments exhibited metal contamination; however, the highest concentrations for most of the metals were measured in the Riá of Huelva sampling sites (MAZ and ML). While TRO had an intermediate metal concentration, SF and the reference sediment (RSP) exhibited the lowest concentrations for most of the metals. The metal concentrations in clear seawater (SWRSP) used for filling the vessels of the experiments were below the detection limits of the equipment with the exception of iron, which had a concentration of 4.97 μg/ L. The mean values for the carbonate system speciation are presented in SI Table S3. These data refer to the carbon parameters calculated in the overlying seawater. The salinity parameter used was 34, and the temperature was 25 °C. Sediment Toxicity Test. The mortality results calculated after the acute 10-d toxicity test with A. brevicornis exposed to various sediment pH treatments are presented in Figure 1. For the RSP sediment samples, the mortality of amphipods was 5% at pH 8.0 (not CO2-added treatment), which validates the use of this site as a reference sediment. No mortality was observed in organisms exposed to SF pH 8.0, and 8% of mortality was detected at TRO pH 8.0. However, more than 70% mortality was observed for the MAZ sediment test at pH 8.0, and approximately 90% mortality was observed in the ML sediment test. Sediments from the Riá of Huelva exhibited metal concentrations over the registered concentrations associated with toxic effects according to site-specific SQG for the Gulf of Cádiz,47 the SQG on the Atlantic Coast of Spain,43 and the 8866



Article

pH. Significant differences were calculated at each pH treatment with respect to the same sediment at its natural pH. Therefore, significant differences (p < 0.01) were observed for the three sediment tests (RSP, SF and MAZ) at pH 7.0, 6.5, and 6.0 compared with the control treatment (pH 8.0). Additionally, mortality measured for the various pH treatments for the studied sediments, SF and TRO, were compared with the mortality at the same pH treatment in the RSP reference sediment. Thus, significant mortality was observed in the TRO sediment test at pH 7.0 (p < 0.05) and pH 7.5 (p < 0.01) compared with the reference sediment. The correlation between mortality and pH decrease can be easily observed in SI Figure S2. A similar mortality trend for the three studied sediments can be observed. The amphipod mortality values calculated after 10 days of exposure time were used to calculate the toxic parameter LpH50. An LpH50 value of 6.86 U was recorded for the RSP sediment test, and LpH50 values of 7.00 U, and 7.20 U were recorded for the SF and TRO sediment tests, respectively. Afterward, the pCO2 threshold that the amphipod organisms could support was calculated by studying pH reductions. Consequently, the RSP pCO2 calculated for pH 6.86 was 13915.1 μatm. For sediment from the SF site, a value of 7.00 was calculated for LpH50, which corresponds to 10 090.0 μatm of pCO2. Finally, a value of LpH50 of 7.20 was calculated for tests using sediment from TRO, meaning a pCO2 of 6301.1 μatm. Carbon dioxide is a toxic substance at elevated concentrations.49 Therefore, there is a pH threshold (or pCO2) that marine organisms can support. If there is a pH reduction below this pH value, then lethal effects on organisms appear. In this study, the pCO2 threshold that amphipod organisms can support was calculated by studying pH reductions. In addition to the acute toxicity associated with CO2-induced acidification, the

Figure 1. Average and standard deviation of amphipod mortality calculated after 10 days of exposure to five sediment toxicity tests subjected to several pH treatments; pH 8.0 or pH control, which means the natural pH of the sediment, pH 7.5 (pH1), pH 7.0 (pH2), pH 6.5 (pH3), and pH 6.0 (pH4). Sediments from SF and TRO were not subjected to pH 6.0 due to the 100% mortality observed for the pH 6.5 treatment. Asterisks indicate significant differences regarding the same sediment at the control pH treatment (*p < 0.05). Common letters indicate significant differences concerning the reference sediment at the same pH treatment (a, p < 0.05; b, P < 0.01).

Spanish Action levels for Dredged Material Management.48 Consequently, the high amphipod mortality percentages presented in MAZ and ML sediment-seawater tests were linked directly to the metal concentration. A mortality rate of 100% was calculated in all the sediment tests for 6.5 or below pH treatments. The studied sediments SF and TRO, as well as the reference sediment, RSP, exhibited mortality rates that were significantly correlated (p < 0.01) with

Figure 2. Metal concentrations in the overlying seawater after 10 days of experiments. The data are the mean metal concentration (μg/L) ± SE (n = 2) at the different pH treatments. The nominal pH treatments are 8.0 (natural sediment pH), 7.5 (treatment 1), 7.0 (treatment 2), 6.5 (treatment 3), and 6.0 (treatment 5). The horizontal line shows USEPA acute (solid line) and chronic (dashed line) recommended seawater quality criteria. Only the RSP sediment was exposed to pH 6.0. Asterisks indicate significant differences regarding the same sediment for the control treatment (*p < 0.05, **p < 0.01). Common letters indicate significant differences with regard to the reference sediment at the same pH treatment (a, p < 0.05; b, p < 0.01). 8867



Article

sediments, RSP and TRO, no significant differences of Cu concentration among the acidification treatments were observed. These results suggest that acidification processes can increase the concentration of dissolved metals in the overlying water, which could affect the toxicity of the sediment to the exposed organisms. Although Cu is an essential micronutrient needed for biochemical processes, it becomes toxic at elevated concentrations.57 Furthermore, Cu is one of the most ubiquitous and pervasive contaminants in marine systems.58 The USEPA recommended seawater quality criteria based the acute value for Cu in 4.8 μg/L.59 Our results indicated that the amount of dissolved Cu was above this quality criteria value at pH 7.0 or less, in sediment from SF. However, amphipod mortality at SF sediment was not significantly different from the reference sediment (RSP) at any pH tested; therefore toxicity is probably attributed to acidification rather than the presence of Cu. The low mobility of Cu observed in the RSP and TRO sediments could be explained due to the metal speciation in the sediments. The chemical form in which the metal is present in the sediment is a key issue determining its mobility and bioavailability. De Orte et al. studied the speciation of metal at the Rio San Pedro area using a sequential extraction technique (BCR) and revealed that 70% of the Cu measured in this sediment was present in the residual fraction of sediments, that means the fraction associated with minerals.45 Metals in this fraction are not likely to be release to the water column, since they are trapped within the crystal structure of the minerals, remaining unavailable to the aquatic system. This could be the reason for the low Cu mobility in the RSP. Regarding TRO, the low Cu mobility could be related to high proportions of organic matter in this site (SI Table S2). Great amounts of Cu in the sediment fraction associated with OM have been reported in several studies of metal speciation.60−62 This is probably due to the high affinity of this metal to humic substances, which are a portion of organic matter. The metals in TRO sediment are likely bound to OM. It could also be adsorbed to hydroxyl groups such as FeOOH, considering the relatively high amount of Fe measured in TRO sediments.63 The mobility of Ni from the sediment to the overlying water increased when the pH decreased in the three study sites (RSP, SF, and TRO). This increase was especially observed in the RSP site, where the differences in Ni concentration between control treatment and the minimum pH tested (pH 6.0) was highly significant (p < 0.01). However, at the TRO test site, Ni amounts at the minimum pH tested (pH 6.5) was also significantly different from the control pH (pH 8.0). In addition, Ni concentration at the overlying water from the TRO site was the highest between the three tested sites, at pH 7.5 and 8.0, for example, the concentration of this metal was significantly different from the Ni measured at RSP site. This is not surprising since the TRO sediment presented higher amounts of this metal than RSP and SF. The concentration of dissolved nickel for all the treatments and sediment tests were always below the USEPA acute quality criteria (74 μg/L); however, the chronic quality value (8.2 μg/L) was exceeded at the lowest pH tested for the TRO (pH 6.5) and the RSP (pH 6.0) sediment tests. This metal exhibited a correlation with pH for the RSP station test. Therefore, Ni could be contributing to amphipod toxicity at this sediment treatment, however, since at this level of acidification, the mortality of the organisms is already 100%, it is not possible to measure the Ni contribution to the toxic responses.

sediment toxicity subjected to CO2 related acidification would be affected by the sediment characteristics. The presented results indicated that depending on the sediment−seawater properties, such physical−chemical characteristics as well as contaminant concentrations, different CO2 enrichments would cause lethal effects to the organisms. Thus, a sediment presenting a previously demonstrated moderate toxicity on estuarine organisms, as TRO,23,50 would require less CO2 injection to cause acute effects on amphipods exposed compared with nontoxic sediments for the organisms, such as the RSP sediment site. The work by Roberts et al. showed a similar response for the amphipod Corophium volutator.51 This study demonstrated that exposure to acidification resulted in greater toxicity of contaminated sediment compared with relatively uncontaminated sediment. In addition, ocean acidification treatments themselves were shown to result in measurable DNA damage in exposed organisms. Hence, a greater susceptibility to lethal and sublethal effects was demonstrated when organisms were exposed to contaminated sediment and elevated pCO2. Nevertheless, this work focused on the near-future ocean acidification (390−1140 μatm pCO2), and the present work attempts not only to predict pCO2 concentrations as the consequence of atmospheric carbon exchanges but, in addition, a much higher pCO2 concentration predicted from potential leaks from subseabed geological storages. Besides, a CO2 leak could be considered as a local problem and a relatively rapid environmental recovery is expected once the CO2 escape is controlled, if more vulnerable habitats, such as coral reef, bed of mussels, and nursery ground areas, were affected by the acidification episode, much time would be needed to completely recuperate the ecosystem. Therefore, the local problem could become larger concerning the environment at large-scale.52 There is lack of information about real CO2 leakages from commercial CO2 storage sites. Therefore, the location and magnitude of a leak is unpredictable.53 Additionally, abrupt sudden releases of CO2 are unlikely but not completely rejected, thus these kinds of events should also be assessed.11,54 Thus, the lowest pH-level presented in this work was selected taking into account: (1) the expected pH threshold that amphipods organisms exposed could support (pH 6.5−6.0); (2) the pH registered in analogous natural CO2 vents, such as that in the Volcanic Ischia Island in Italy (lowest pH registered 6.57),55 and the natural CO2 vent recently discovered above a natural salt dome at the Southern North Sea (pH 6.8 ± 0.2) ;56 and (3) considering that, even small but continuous seepages of CO2 may be sufficient to cause major reductions of pH (pH 6.5) in the sediment−water interface.14 Effect of Acidification on Metal Mobility. The effect of acidification on metal mobilization from the sediment to the test overlying seawater is depicted in Figure 2. Among all the metals measured, Cu, Ni, and Zn were mobilized from the sediment to the overlying water during the RSP, SF, or TRO sediment tests due to pH reductions. Regarding the metal Cu, its concentration increased due to pH reductions, in overlying water of the SF sediment toxicity test. This was mainly observed at the lowest pH treatment tested, pH 6.5, where a significant difference (p < 0.05) in the concentration of this metal was observed when compared to the control pH treatment (pH 8.0). In addition, the amount of Cu in the overlying seawater during the SF sediment test at pH 6.5 presented a significant difference (p < 0.05) with the RSP sediment test at the same pH treatment. About the other test 8868



Article

FeOOH, perhaps aggregated to organic matter could be explaining the immobility of the metal in these sediments.63,68 In order to clarify the link between toxicity results and metal mobilization, a principal component analysis (PCA) was preformed (SI Table S4). Two new factors were obtained from the original values. These factors, taken together, explained almost 80% of the total variance. Factor 1 accounted for 55% of the total variance and correlated the amphipods mortality, the metals Ni and Zn, and the pCO2 with the pH, through negative values. This analysis suggests that the dissolved metals (Ni and Zn) are linked to acidification and toxicity, however further research on the effects of CO2-related acidification on metal mobility and linked toxicity are needed to conclude this relationship. The concentrations of the metals Cr, Cd, and As were below the detection limits of the equipment in the overlying water of all study sites. Finally, an increase of the Co and Pb concentrations in the RSP and TRO sediment tests could be observed for the lowest pH treatments, however this increase was not statistically relevant. When assessing the mobility of metals in a sediment-water environment, it should be considered that the mobility of the metals present in marine sediments depends on their association with the solid phases to which they are bound.70 The acidification of the samples tends to release the metals that are less strongly associated with sediments increasing their potential bioavailability.27 In this work, we could observe how sediment composition influences metal mobilization. For example, in TRO sediment, the high amounts of OM were probably the reason for lower Cu mobilization when compared to the other study sites, even though total Cu concentration was higher at TRO sediment. In this sense, experimental work regarding CCS risks should consider using sediments with different characteristics (mineralogy, granulometry, etc.) in order to have a better understanding on the consequences of CO2 leakage scenarios. Previous studies have reported the bioaccumulation of metals in bivalves due to acidification.64,35 However, the authors manipulated the pH through adjustments with HCl and NaOH. Although this method has been contemplated in the “Guide to Best Practices for Ocean Acidification Research and Data Reporting”,71 it has been demonstrated that this procedure less realistically simulates CO2 leaks because it modifies the total alkalinity instead of increasing the concentration of dissolved inorganic carbon.72,73 Furthermore, studies that compared the effect of acidification by a strong acid and by CO2 in the early stages of sea bream and on the growth of the microalgae Phaeodactylum tricornutum noted that acidification by an acid could lead to underestimation of the carbonate chemistry alteration expected by CO2 acidification.74 To date, most of the works assessing the risk associated with CCS technologies have been focused on geochemical alteration or contaminant availability due to acidification processes.67,12,14,45,75 However, little is known about the implication of CO2 leaks on the marine biota.66,35,44,76,77,72 Metals are frequent contaminants in sediment, and their bioavailability is an important tool to understand the hazard that represents on the organism exposed.78 Furthermore, it is known that metals are largely affected by acidification processes. Thus, studies addressing the environment implications of CCS technologies are needed. Beside of the risk associated with CCS technologies, it is an acknowledged reality that the oceans are becoming more acidic because of the CO2 atmosphere−ocean exchanges.79 Both

Within the toxic metals, Zn presented the highest concentrations in the overlying water. The mobility of this metal from the sediment to the overlying seawater had a similar pattern in the three tested sites, showing increases in Zn concentration following pH decreases. The concentration of Zn was significantly different from the control pH (pH 8.0) even at intermediate pH treatments (pH 7.0) for both RSP (p < 0.01) and SF (p < 0.05) study sites. Despite Zn amounts being increased in all sites, its concentration at SF and TRO were significantly higher (p < 0.05) than the reference sediment (RSP) at all tested pH treatments. It should be noted that the Zn concentration in the whole sediment was approximately 3-fold higher for the sediment from TRO (156 mg/kg) than for those from SF (53.8 mg/kg) and RSP (49 mg/kg). The higher amounts of Zn at the TRO site could be the reason for the greater mortality rates in amphipods at intermediate pH levels (pH 7.5 and 7.0) when compared to RSP and SF toxic responses. According to acute toxic values from USEPA recommended seawater quality criteria for Zn (90 μg/L), concentration of this metal would be lethal at the lowest pH treatments for both TRO and SF. However, combination of two toxic parameters, pH and the presence of Zn, could be increasing the toxicity at the intermediate pHs. Similar results associating the increase in toxic effects on bivalve organisms to metal mobilization, especially Zn, from sediment to water were reported by Riba et al.64 Contrasting results were observed by Roberts et al., who demonstrated that the flux of Zn from the sediment to the overlaying water was unaffected by pCO2 treatments and that significant differences were not observed between a contaminated sediment and the reference sediment.51 With regard to Fe, mobility from sediment to the overlying water did not follow any specific pattern related to acidification. No significant differences in Fe concentration were observed between pH treatments. Additionally, differences in Fe concentration between study sites were not significant either. The iron solubility in surface seawater is low. It is known, for example, that the lifetime of dissolved iron depends on its oxidation rate, which is strongly dependent on pH.65 It is expected that pH decreases in the seawater will lead to increases in the half-life of dissolved Fe. In this work, measurements of this metal were only performed at the end of the experiment. The concentration of Fe in the overlying water may have increased due to acidification processes at the beginning of the experiment, however, the rapid oxidative precipitation of this metal could be responsible for Fe precipitation during experimental time and for this reason differences in the amount of this metal between pH treatments could not be observed. This behavior was observed before by De Orte et al. that reported increases in Fe concentration due to acidification after 24 h exposures in CO2 leakage experiments. However, after 10 days of experiment, Fe concentrations decreased greatly in all pH treatments tested.66 Ardeland and Steinnes reported the mobilization of Fe DGT labile fraction from sediment to overlying seawater due to CO2 seepages.67 This mobilization was also affected by sediment resuspension caused by gas seepages. Furthermore, the authors concluded a disturbance of Fe−Mn shuttle in the sediment, which could lead to enhanced concentrations of toxic metals and trace elements in seawater. The sediments, RSP and TRO, present a relatively high Fe concentration, and a similar behavior to Cu as to pH could be observed. Because, among the particulates capable of absorbing Cu are the hydrous ferric oxides,68,69 the adsorption of Cu to 8869



■

ACKNOWLEDGMENTS Research presented in this document was supported partially by the Spanish Ministerio de Economiá y Competitividad, under Grant Reference CTM2011-28437-C02-02/TECNO and CTM2012-36476-C02-01/TECNO. The authors are also grateful to international Grant from Bank Santander/UNESCO Chair UNITWIN/WiCop for funding this work. M.D.B. thanks A. Rodriguez-Romero and Dr. J. Blasco for their valuable collaboration on the performance of the experiments and for their personal clarifications.We also thank the reviewers of the American Journal Expert for correcting the language, grammar, and style. The authors also would like to give special acknowledgements to the four anonymous reviewers for their valuable comments on this manuscript.

acidification processes, CO2 leaks and ocean acidification, could lead to perturbations in marine ecosystems, which need to be investigated widely. The presented results suggest that acidification processes can increase the concentration of dissolved metals in the overlying water, which could affect the toxicity of the sediment to the exposed organism. Because coastal and estuarine areas are frequently contaminated by metals, the quality of the sediments could be affected by pH decreases leading to increases in toxicity of sediments that are not presenting toxic effects at natural ambient pH, i.e., by increasing metal bioavailability. This acidification process must be considered when evaluating the quality of sediments by sediment guidelines, especially in heavily metal contaminated areas such as the Riá of Huelva estuary. According to our results, the expected pH reduction for the near-future (up 0.5 U by 2100)80,79,81 will not cause lethal effects to the amphipods exposed and neither to metal mobilization. Furthermore, adaptation capacity of marine organisms to slow pH variations has been previously reported.6,82−85 Nonetheless, taking into account the future perspective related to CCS projects (12 projects in operations, 9 under construction, and 39 in various stages of development planning) it is considered essential to evaluate the potential effects of this technology in the environment, especially taking into account the uncertainties related to the reliability of this activity.

■

■

ABBREVIATIONS CCS Carbon Capture and Storage RSP Rió San Pedro SF San Fernando TRO Trocadero MAZ Mazagón ML Muelle de Levante

■

REFERENCES

(1) UNFCCC, United Nations Framework Convention on Climate Change. United Nations, FCCC/INFORMAL/84 GE.05−62220 (E) 200705.: 1992 http://unfccc.int/resource/docs/convkp/conveng.pdf. (2) London Protocol, Specific guidelines for the assessment of carbon dioxide streams for disposal into sub-seabed geological formations. 1996 London protocol on the prevention of marine pollution by dumping of wastes and other matter. 2007 http://www.gc.noaa.gov/documents/ gcil_imo_co2wag.pdf. (3) OSPAR Convention, Guidelines for risk assessment and management of storage of carbon dioxide streams in sub-seabed geological formations. OSPAR Convention for the protection of the marine environment of the North-East Atlantic. 2007. (4) Goldberg, D. S.; Takahashi, T.; Slagle, A. L. Carbon dioxide sequestration in deep-sea basalt. Proc. Natl. Acad. Sci. U. S. A. 2008, 105 (29), 9920−9925. (5) Hawkins, D. G. No exit: Thinking about leakage from geologic carbon storage sites. Energy 2004, 29 (9−10), 1571−1578. (6) Global CCS Institute. The Global Status of CCS: Project 2014; Canberra, Australia, 2014 http://cdn.globalccsinstitute.com/sites/ default/files/publications/22562/global-status-ccs-2011.pdf. (7) (a) BOE, Resolution of 28 of November 2007. Ministry of Industry, Tourism and Trade: 2008; Vol. 34, pp 7099−7102. (b) BOE. Resolution of 4 of March of 2008. Ministry of Industry. Tourism Trade 2008; Vol. 81, p 18586. (8) Blackford, J.; Jones, N.; Proctor, R.; Holt, J.; Widdicombe, S.; Lowe, D.; Rees, A. An initial assessment of the potential environmental impact of CO2 escape from marine carbon capture and storage systems. Proc. Inst. Mech. Eng. A 2009, 223 (3), 269−280. (9) Dethlefsen, F.; Köber, R.; Schäfer, D.; Hagrey, S. A. a.; Hornbruch, G.; Ebert, M.; Beyer, M.; Großmann, J.; Dahmke, A. Monitoring approaches for detecting and evaluating CO2 and formation water leakages into near-surface aquifers. Energy Procedia 2013, 37 (0), 4886− 4893. (10) Koornneef, J.; Spruijt, M.; Molag, M.; Ramírez, A.; Turkenburg, W.; Faaij, A. Quantitative risk assessment of CO2 transport by pipelinesA review of uncertainties and their impacts. J. Hazard. Mater. 2010, 177 (1−3), 12−27. (11) Bruant, R. G., Jr; Celia, M. A.; Guswa, A. J.; Peters, C. A. Peer reviewed: Safe storage of CO2 in deep saline aquifiers. Environ. Sci. Technol. 2002, 36 (11), 240A−245A. (12) Ardelan, M. V.; Steinnes, E.; Lierhagen, S.; Linde, S. O. Effects of experimental CO2 leakage on solubility and transport of seven trace

ASSOCIATED CONTENT

S Supporting Information *

Figure S1 presents the map of the location of the water and sediment collection showing the general areas of the samples sites. Figure S2 is showing the mean values of LpH50 calculated for the A. brevicronis organisms at the three sediment sites RSP, SF, and TRO. Tables S1 and S2 provide a summarized description of test parameters and conditions for the amphipod toxicity tests and the different pH treatments and the carbonate system speciation in sediment-seawater experiments at each pH treatment, respectively. Table S3 provides the carbonate system speciation in sediment-seawater experiments at each pH treatment. Table S4 presents the rotated component matrix with component loadings of the original variables for the two principal factors after multivariate analysis of the result obtained. This material is available free of charge via the Internet at http:// pubs.acs.org/.

■

Article

AUTHOR INFORMATION

Corresponding Author

*Tel: +34 956017334; fax: +34 956016040; e-mail: dolores. [email protected]. Present Address ‡

Instituto do Mar, Campus Baixada Santista, Universidade Federal de São Paulo, Av. Alm. Sandanha da Gama, 89-Ponta da Praia/SP CEP: 11030−400. Santos, SP, Brazil. Tel: +55(13) 35235094. Author Contributions

The manuscript was written through contributions of all authors. All authors have given approval to the final version of the manuscript. Notes

The authors declare no competing financial interest. 8870



Article

metals in seawater and sediment. Sci. Total Environ. 2009, 407 (24), 6255−6266. (13) Widdicombe, S.; Beesley, A.; Berge, J. A.; Dashfield, S. L.; McNeill, C. L.; Needham, H. R.; Øxnevad, S. Impact of elevated levels of CO2 on animal mediated ecosystem function: The modification of sediment nutrient fluxes by burrowing urchins. Mar. Pollut. Bull. 2013, 73 (2), 416−427. (14) Ardelan, M. V.; Sundeng, K.; Slinde, G. A.; Gjøsund, N. S.; Nordtug, T.; Olsen, A. J.; Steinnes, E.; Torp, T. A. Impacts of Possible CO2 Seepage from Sub-Seabed Storage on Trace Elements Mobility and Bacterial Distribution at Sediment-Water Interface. Energy Procedia 2012, 23 (0), 449−461. (15) Rochelle, C.; Czernichowski-Lauriol, I.; Milodowski, A. The impact of chemical reactions on CO2 storage in geological formations: A brief review. Geological Society; Special Publications, London 2004, 233 (1), 87−106. (16) Carroll, A. G.; Przeslawski, R.; Radke, L. C.; Black, J. R.; Picard, K.; Moreau, J. W.; Haese, R. R.; Nichol, S. Environmental considerations for subseabed geological storage of CO2: A review. Cont. Shelf Res. 2014, 83 (0), 116−128. (17) Conradi, M.; Depledge, M. H. Population responses of the marine amphipod Corophium volutator (Pallas, 1766) to copper. Aquat. Toxicol. 1998, 44 (1), 31−45. (18) ASTM. Standard guide for conducting 10-day static sediment toxicity tests with marine and estuarine amphipods. In Publ. E., Philadelphia, 1993; pp 1367−1392, 26pp. (19) Environment Canada, Biological Test Method: Acute Test for Sediment Toxicity Using Marine or Estuarine Amphipods. In Report EPS 1/RM/26, Environmental Protection, Conservation and Protection: Ottawa, Ontario, 1992. (20) USEPA. Methods for assessing the toxicity of sediment-associated contaminants with estuarine and marine amphipods. United State Environmental Protection Agency 1994; Vol. EPA/6007R-94/025. (21) Casado-Martínez, M. C.; Forja, J. M.; DelValls, T. A. Direct comparison of amphipod sensitivities to dredged sediments from Spanish ports. Chemosphere 2007, 68 (4), 677−685. (22) Riba, I.; DelValls, T. A.; Forja, J.; Gómez-Parra, A. Comparative toxicity of contaminated sediment from a mining spill using two amphipods species: Corophium volutator (Pallas, 1776) and Ampelisca brevicornis (A. Costa, 1853). Bull. Environ. Contam. Toxicol. 2003, 71 (5), 1061−1068. (23) DelValls, T. Á .; Forja, J. M.; Gómez-Parra, A. Integrative assessment of sediment quality in two littoral ecosystems from the Gulf of Cádiz, Spain. Environ. Toxicol. Chem. 1998, 17 (6), 1073−1084. (24) Ligero, R. A.; Barrera, M.; Casas-Ruiz, M.; Sales, D.; LópezAguayo, F. Dating of marine sediments and time evolution of heavy metal concentrations in the Bay of Cãdiz, Spain. Environ. Pollut. 2002, 118 (1), 97−108. (25) Silva, C.; Yáñez, E.; Martín-Díaz, M.; DelValls, T. Assessing a bioremediation strategy in a shallow coastal system affected by a fish farm culture − Application of GIS and shellfish dynamic models in the Rio San Pedro, SW Spain. Mar. Pollut. Bull. 2012, 64, 751−765. (26) Borrego, J.; Morales, J.; de la Torre, M.; Grande, J. Geochemical characteristics of heavy metal pollution in surface sediments of the Tinto and Odiel river estuary (southwestern Spain). Environ. Geol. 2002, 41 (7), 785−796. (27) Riba, I.; Delvalls, T. Á .; Forja, J. M.; Gómez-Parra, A. The influence of pH and salinity on the toxicity of heavy metals in sediment to the estuarine clam Ruditapes philippinarum. Environ. Toxicol. Chem. 2004, 23 (5), 1100−1107. (28) Riba, I.; Forja, J. M.; Gómez-Parra, A.; DelValls, T. Á . Sediment quality in littoral regions of the Gulf of Cádiz: A triad approach to address the influence of mining activities. Environ. Pollut. 2004, 132 (2), 341−353. (29) Gaudette, H. E.; Flight, W. R.; Toner, L.; Folger, D. W. An inexpensive titration method for the determination of organic carbon in recent sediments. J. Sediment Petrol 1974, 44 (1), 249−53.

(30) El-Rayis, Re-assessment of the titration method for determination of organic carbon in recent sediments. Rapp Comm Int. Mer Médit 1985, 29, 45−7. (31) ASTMD422-63, Standard test method for particle-size analysis of soils. 2007. (32) Gee, G. W.; Or, D. 2.4 Particle-size analysis. Methods Soil Anal. 2002, 4, 255−293. (33) Flemming, B. A revised textural classification of gravel-free muddy sediments on the basis of ternary diagrams. Cont. Shelf Res. 2000, 20 (10), 1125−1137. (34) Loring, D.; Rantala, R. Manual for the geochemical analyses of marine sediments and suspended particulate matter. Earth-Sci. Rev. 1992, 32 (4), 235−283. (35) Riba, I.; Kalman, J.; Vale, C.; Blasco, J. Influence of sediment acidification on the bioaccumulation of metals in Ruditapes philippinarum. Environ. Sci. Pollut. Res. 2010, 17 (9), 1519−1528. (36) Pierrot, D.; Lewis, E.; RWallace, D. W. CO2SYS Dos program developed for CO2 system calculations. In ORNL/CDIAC-105. Carbon Dioxide Information Analysis Center. Department of Energy; Oak Ridge National Laboratory: Oak Ridge, Tennessee, 2006. (37) Mehrbach, C.; Culberson, C.; Hawley, J.; Pytkowicz, R. Measurement of the apparent dissociation constants of carbonic acid in seawater at atmospheric pressure. Limnol. Oceanogr. 1973, 897−907. (38) Dickson, A. G.; Millero, F. J. A comparison of the equilibrium constants for the dissociation of carbonic acid in seawater media. Deep Sea Res. A. 1987, 34 (10), 1733−1743. (39) Dickson, A. G. Standard potential of the (AgCl(s) + 1/2H2 (g) = Ag(s) + HCl(aq)) cell and the dissociation constant of bisulfate ion in synthetic sea water from 273.15 to 318.15 K. J. Chem. Thermodyn. 1990, 22, 113−123. (40) Casado-Martínez, M. C.; Buceta, J. L.; Belzunce, M. J.; DelValls, T. A. Using sediment quality guidelines for dredged material management in commercial ports from Spain. Environ. Int. 2006, 32 (3), 388−396. (41) Morales-Caselles, C.; Kalman, J.; Riba, I.; DelValls, T. A. Comparing sediment quality in Spanish littoral areas affected by acute (Prestige, 2002) and chronic (Bay of Algeciras) oil spills. Environ. Pollut. 2007, 146 (1), 233−240. (42) Ramos-Gómez, J.; Martín-Díaz, M.; DelValls, T. Acute toxicity measured in the amphipod Ampelisca brevicornis after exposure to contaminated sediments from Spanish littoral. Ecotoxicology 2009, 18 (8), 1068−1076. (43) Riba, I.; Casado-Martínez, C.; Forja, J. M.; DelValls, Á . Sediment quality in the Atlantic coast of Spain. Environ. Toxicol. Chem. 2004, 23 (2), 271−282. (44) Basallote, M.; Rodríguez-Romero, A.; Blasco, J.; DelValls, A.; Riba, I. Lethal effects on different marine organisms, associated with sediment−seawater acidification deriving from CO2 leakage. Environ. Sci. Pollut. Res. 2012, 19 (7), 2550−2560. (45) De Orte, M. R.; Sarmiento, A. M.; Basallote, M. D.; RodríguezRomero, A.; Riba, I.; delValls, A. Effects on the mobility of metals from acidification caused by possible CO2 leakage from sub-seabed geological formations. Sci. Total Environ. 2014, 470−471 (0), 356−363. (46) DelValls, T.; Casado-Martínez, M.; Riba, I.; Martín-Díaz, M.; Forja, J.; García-Luque, E.; Gómez-Parra, A.Technical Report for CEDEX: Investigación conjunta sobre la viabilidad de utilizar ensayos ecotoxicológicos para la evaluación de la calidad ambiental del material de dragado. Puerto Real (Cádiz), 2003 (47) Choueri, R. B.; Cesar, A.; Abessa, D. M. S.; Torres, R. J.; Morais, R. D.; Riba, I.; Pereira, C. D. S.; Nascimento, M. R. L.; Mozeto, A. A.; DelValls, T. A. Development of site-specific sediment quality guidelines for North and South Atlantic littoral zones: Comparison against national and international sediment quality benchmarks. J. Hazard. Mater. 2009, 170 (1), 320−331. (48) CEDEX, Spanish Action Levels for dredged material management. Recomendation for the management of dredged material in the ports of Spain. C. d. e. y. e. d. o. públicas, Ed. Puertos del Estado. Madrid., 1994. 8871



Article

(49) Hsieh, C.-C.; Shih, C.-L.; Fang, C.-C.; Chen, W.-J.; Lee, C.-C. Carbon dioxide asphyxiation caused by special-effect dry ice in an election campaign. Am. J. Emerg. Med. 2005, 23 (4), 567−568. (50) Araújo, C. V. M; Diz, F. R.; Laiz, I.; Lubian, L. M.; Blasco, J.; Moreno-Garrido, I. Sediment integrative assessment of the Bay of Cádiz (Spain): An ecotoxicological and chemical approach. Environ. Int. 2009, 35 (6), 831−841. (51) Roberts, D. A.; Birchenough, S. N.; Lewis, C.; Sanders, M. B.; Bolam, T.; Sheahan, D. Ocean acidification increases the toxicity of contaminated sediments. Global Change Biol. 2013, 19 (2), 340−351. (52) http://www.eco2-project.eu/. (53) Dewar, M.; Wei, W.; McNeil, D.; Chen, B. Small-scale modelling of the physiochemical impacts of CO2 leaked from sub-seabed reservoirs or pipelines within the North Sea and surrounding waters. Mar. Pollut. Bull. 2013, 73 (2), 504−515. (54) Lewicki, J. L.; Oldenburg, C. M.; Dobeck, L.; Spangler, L. Surface CO2 leakage during two shallow subsurface CO2 releases. Geophys. Res. Lett. 2007, 34 (24), L24402. (55) Hall-Spencer, J. M.; Rodolfo-Metalpa, R.; Martin, S.; Ransome, E.; Fine, M.; Turner, S. M.; Rowley, S. J.; Tedesco, D.; Buia, M.-C. Volcanic carbon dioxide vents show ecosystem effects of ocean acidification. Nature 2008, 454 (7200), 96−99. (56) McGinnis, D. F.; Schmidt, M.; DelSontro, T.; Themann, S.; Rovelli, L.; Reitz, A.; Linke, P. Discovery of a natural CO2 seep in the German North Sea: Implications for shallow dissolved gas and seep detection. J. Geophys. Res.: Oceans 2011, 116 (C3), C03013. (57) Hebel, D.; Jones, M.; Depledge, M. Responses of crustaceans to contaminant exposure: a holistic approach. Estuarine, Coastal Shelf Sci. 1997, 44 (2), 177−184. (58) Eisler, R. Copper Hazards to Fish, Wildlife, And Invertebrates: A Synoptic Review; US Dept. of Commerce: Springfield, 1998. (59) USEPA http://water.epa.gov/scitech/swguidance/standards/ criteria/current/index.cfm. (60) Li, Q.; Wu, Z.; Chu, B.; Zhang, N.; Cai, S.; Fang, J. Heavy metals in coastal wetland sediments of the Pearl River Estuary, China. Environ. Pollut. 2007, 149 (2), 158−164. (61) Morillo, J.; Usero, J.; Gracia, I. Heavy metal distribution in marine sediments from the southwest coast of Spain. Chemosphere 2004, 55 (3), 431−442. (62) Passos, E. d. A.; Alves, J. C.; dos Santos, I. S.; Alves, J. d. P. H.; Garcia, C. A. B.; Spinola Costa, A. C. Assessment of trace metals contamination in estuarine sediments using a sequential extraction technique and principal component analysis. Microchem. J. 2010, 96 (1), 50−57. (63) Chapman, P. M.; Wang, F.; Janssen, C.; Persoone, G.; Allen, H. E. Ecotoxicology of metals in aquatic sediments: binding and release, bioavailability, risk assessment, and remediation. Can. J. Fish. Aquat. Sci. 1998, 55 (10), 2221−2243. (64) Riba, I.; Garcia-Luque, E.; Blasco, J.; DelValls, T. Bioavailability of heavy metals bound to estuarine sediments as a function of pH and salinity values. Chem. Spec. Bioavail. 2003, 15 (4), 101−114. (65) Millero, F. J.; Sotolongo, S.; Izaguirre, M. The oxidation kinetics of Fe(II) in seawater. Geochim. Cosmochim. Acta 1987, 51 (4), 793−801. (66) De Orte, M. R.; Lombardi, A. T.; Sarmiento, A. M.; Basallote, M. D.; Rodriguez-Romero, A.; Riba, I.; Del Valls, A. Metal mobility and toxicity to microalgae associated with acidification of sediments: CO2 and acid comparison. Mar. Environ. Res. 2014, 96 (0), 136−144. (67) Ardelan, M. V.; Steinnes, E. Changes in mobility and solubility of the redox sensitive metals Fe, Mn, and Co at the seawater-sediment interface following CO2 seepage. Biogeosciences 2010, 7 (2), 569−583. (68) Dzombak, D. A.; Morel, F. M. M. Surface Complexation Modelling: Hydrous Ferric Oxide; New York, 1990. (69) Richards, R.; Chaloupka, M.; Sano, M.; Tomlinson, R. Modelling the effects of “coastal” acidification on copper speciation. Ecol. Model. 2011, 222 (19), 3559−3567. (70) Ure, A.; Davidson, C.Chemical speciation in the environment; 2001. (71) Riebesell, U.; Fabry, V. J.; Hansson, L.; Gattuso, J. P. Guide to best practices for ocean acidification research and data reporting; Publications Office of the European Union: 2010.

(72) Pascal, P.-Y.; Fleeger, J. W.; Galvez, F.; Carman, K. R. The toxicological interaction between ocean acidity and metals in coastal meiobenthic copepods. Mar. Pollut. Bull. 2010, 60 (12), 2201−2208. (73) Hurd, C. L.; Hepburn, C. D.; Currie, K. I.; Raven, J. A.; Hunter, K. A. Testing the effects of ocean acidification on algal metabolism: Considerations for experimental designs. J. Phycol. 2009, 45 (6), 1236− 1251. (74) Kikkawa, T.; Kita, J.; Ishimatsu, A. Comparison of the lethal effect of CO2 and acidification on red sea bream (Pagrus major) during the early developmental stages. Mar. Pollut. Bull. 2004, 48 (1), 108−110. (75) Payan, M. C.; Verbinnen, B.; Galan, B.; Coz, A.; Vandecasteele, C.; Viguri, J. R. Potential influence of CO2 release from a carbon capture storage site on release of trace metals from marine sediment. Environ. Pollut. 2012, 162 (0), 29−39. (76) Rodríguez-Romero, A.; Basallote, M. D.; Orte, M. R. D.; DelValls, T. Á .; Riba, I.; Blasco, J. Simulation of CO2 leakages during injection and storage in sub-seabed geological formations: Metal mobilization and biota effects. Environ. Int. 2014, 68, 105−117. (77) Fitzer, S. C.; Caldwell, G. S.; Clare, A. S.; Upstill-Goddard, R. C.; Bentley, M. G. Response of copepods to elevated pCO2 and environmental copper as co-stressorsA multigenerational study. PloS one 2013, 8 (8), e71257. (78) Martín-Díaz, M. L.; Jiménez-Tenorio, N.; Sales, D.; DelValls, T. A. Accumulation and histopathological damage in the clam Ruditapes philippinarum and the crab Carcinus maenas to assess sediment toxicity in Spanish ports. Chemosphere 2008, 71 (10), 1916−1927. (79) Caldeira, K.; Wickett, M. E. Ocean model predictions of chemistry changes from carbon dioxide emissions to the atmosphere and ocean. J. Geophys. Res. 2005, 110 (C9), C09S04. (80) Caldeira, K.; Wickett, M. E. Oceanography: Anthropogenic carbon and ocean pH. Nature 2003, 425 (6956), 365−365. (81) Blackford, J. C.; Gilbert, F. J. pH variability and CO2 induced acidification in the North Sea. J. Mar. Syst. 2007, 64 (1−4), 229−241. (82) Sunday, J. M.; Crim, R. N.; Harley, C. D. G.; Hart, M. W. Quantifying rates of evolutionary adaptation in response to ocean acidification. PloS One 2011, 6 (8), e22881. (83) Hofmann, G. E.; Todgham, A. E. Living in the now: Physiological mechanisms to tolerate a rapidly changing environment. Ann. Rev. Physiol. 2010, 72, 127−145. (84) Melzner, F.; Gutowska, M.; Langenbuch, M.; Dupont, S.; Lucassen, M.; Thorndyke, M. C.; Bleich, M.; Pö rtner, H. O. Physiological basis for high CO2 tolerance in marine ectothermic animals: Pre-adaptation through lifestyle and ontogeny? Biogeosciences (BG) 2009, 6, 2313−2331. (85) Pistevos, J. C. A.; Calosi, P.; Widdicombe, S.; Bishop, J. D. D. Will variation among genetic individuals influence species responses to global climate change? Oikos 2011, 120 (5), 675−689.

8872


Studying the Effect of CO2-Induced ... - ACS Publications

Recommend Documents