Metal mobility and toxicity to microalgae associated with acidification of sediments: CO2 and acid comparison

Marine Environmental Research 96 (2014) 136e144

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Metal mobility and toxicity to microalgae associated with acidification of sediments: CO2 and acid comparison M.R. De Orte a, *, A.T. Lombardi b, A.M. Sarmiento a, c, M.D. Basallote a, A. Rodriguez-Romero d, I. Riba a, A. Del Valls a a

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 Departamento de Botânica, Universidade Federal de São Carlos, Rodovia Washington Luis Km 235, 13565-905 São Carlos, Brazil c Departamento de Geología, F. Ciencias Experimentales, Campus El Carmen, Universidad de Huelva, Avda. Fuerzas Armadas, s/n, 21071 Huelva, Spain d Departamento de Ecología y Gestión Costera, Instituto de Ciencias Marinas de Andalucía (CSIC), Campus Río San Pedro, 11510 Puerto Real, Cádiz, Spain b

a r t i c l e i n f o

a b s t r a c t

Article history: Received 31 July 2013 Received in revised form 30 September 2013 Accepted 3 October 2013

The injection and storage of CO2 into marine geological formations has been suggested as a mitigation measure to prevent global warming. However, storage leaks are possible resulting in several effects in the ecosystem. Laboratory-scale experiments were performed to evaluate the effects of CO2 leakage on the fate of metals and on the growth of the microalgae Phaeodactylum tricornutum. Metal contaminated sediments were collected and submitted to acidification by means of CO2 injection or by adding HCl. Sediments elutriate were prepared to perform toxicity tests. The results showed that sediment acidification enhanced the release of metals to elutriates. Iron and zinc were the metals most influenced by this process and their concentration increased greatly with pH decreases. Diatom growth was inhibited by both processes: acidification and the presence of metals. Data obtained is this study is useful to calculate the potential risk of CCS activities to the marine environment. Ó 2013 Elsevier Ltd. All rights reserved.

Keywords: Carbon capture and storage (CCS) Acidification Metal mobility Microalgae Toxicity

1. Introduction The concentration of CO2 in the atmosphere rose throughout the twentieth century and keeps rising mainly as a result of energy production, followed by land use practices, such as deforestation (IPCC, 2007). The linkage between increases of this greenhouse gas and climate change phenomena has lead to the concern of mitigation measures to reduce the emission of this gas, such as moving to renewable energy generation, increasing energy efficiency, better land use practices, bio-energy, shifting to more nuclear power, etc. Among the options, large-scale CO2 capture and storage in marine geological formation (CCS) is under study. Nowadays there are several ongoing CCS projects and general scientific agreement on the economics and environmental viability of them (Reguera et al., 2009). Nevertheless, both knowledge and experience in this type of activity is still under development, especially regarding the environmental risks involved. One of the main risks associated to CCS is the leakage of the stored CO2 and the consequent acidification of the marine and estuarine environment

* Corresponding author. Tel./fax: þ34 956016040. E-mail address: [email protected] (M.R. De Orte). 0141-1136/$ e see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.marenvres.2013.10.003

(Ardelan et al., 2009; Ardelan and Steinnes, 2010; Blackford et al., 2009). Decreases in seawater pH due to increase in CO2 concentration may result in potential impacts in marine ecosystems, affecting marine organisms and community structure (Basallote et al., 2012; Ishimatsu et al., 2004; Hale et al., 2011; Kikkawa et al., 2006; Widdicombe et al., 2009) by interfering in physiological processes such as growth, development, metabolism, calcification, ion regulation and acidebase balance. In addition to biological changes, geochemical alterations are also expected, such as the mobilization of metals previously trapped in sediments (Ardelan et al., 2009; Ardelan and Steinnes, 2010; De Orte et al., 2013; Payán et al., 2012). Moreover, the acidification of the ocean will alter the organic and inorganic speciation of metals and will modify their interaction with organisms (Millero et al., 2009). Due to the relevance of this issue and its global nature, there are several renowned research groups evaluating the effects of CO2 leakage through in situ experiments and laboratory simulations (Ardelan et al., 2009; Ardelan and Steinnes, 2010; Basallote et al., 2012; De Orte et al., 2013; Keating et al., 2011; Payán et al., 2012; Widdicombe et al., 2009). At the same time that risks of CCS activities and establishment of appropriate risk management are evaluated, remediation plans are being investigated. However, until

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Fig. 1. Map of the two study areas, Huelva estuary, showing sampling stations MZ and LEV, and the Bay of Cádiz, showing sampling station RSP, all located in the south-western of the Iberian Peninsula.

now no acceptable CO2 leak has been defined, and at the same time, the limits of effectiveness of geological storage as a measure to reduce CO2 emissions and mitigate climate change in the long term has not been recognized. As the major primary producer, phytoplankton play essential role in the marine environment, and any interference in their community structure could end up affecting the whole ecosystem. Furthermore, since they are in direct contact with the medium, only been separated by a cell wall and membrane, they are highly affected by the presence of pollutants and therefore they have been used as biological indicators to measure pollutant toxicity (Franqueira et al., 2000). It is known that increase in the concentration of CO2 can affect the growth rate of microalgae population, altering microalgae community assemblages (Tortell et al., 2002). Nevertheless studies regarding the interferences of CO2 enrichment in microalgae are limited to those scenarios mimicking ocean acidification due to the uptake of CO2 (with the lowest pH around 7.8), and as far as we are concerned, there are no studies on microalgae responses, regarding CO2 leakage scenarios using pH values lower than natural seawater pH. The first aim of this work was to study the effect of CO2 leakages from CCS activities on the release and geochemistry of metals from

contaminated marine sediment under different levels of acidification. The second aim of this study was to evaluate the toxicity resulting from the acidification and its relations with metal release to the marine diatom Phaeodactylum tricornutum by comparing two different methodologies for acidification: CO2 bubbling and HCl addition. Laboratory controlled microalgae cultures were performed using natural elutriates as growth medium. A CO2 bubbling system was designed to conduct short-term toxicity tests under laboratory conditions (non-pressurized) employing a range of pH treatments. The effects of elevated CO2 levels on the microalgae were studied using standardized toxicity test guidelines. Experiments were performed using sediment samples collected in Huelva estuary, which is located on the Atlantic coast of south-western Spain. This estuary was chosen as study area because it is heavily contaminated by metals. It has different sources of contamination including industrial residues from chemical industry plants located in areas close to the Ría of Huelva, the urban sewage from the city of Huelva, and, most importantly, the fluvial inputs from Tinto and Odiel rivers, which, due to a long period of intensive mining activities within their basins, contain high concentrations of toxic metals (Sarmiento et al., 2011) and residues from acid mine drainage (Sarmiento

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et al., 2009a, 2009b). In addition, another area was selected as a control or non-contaminated site, the shallow tidal creek area of the Río San Pedro located some distance to the southeast of the Huelva estuary, in the Bay of Cádiz, sediments in this area contain low levels of metals (Blasco, 2005). 2. Materials and methods 2.1. Sampling Two different sampling sites were selected in the Huelva estuary. One is located in the inner part of the estuary within the Odiel River itself (LEV) and the other is located at the mouth of the estuary (MZ) (Fig. 1). In the Río San Pedro only one site (RSP) (Fig. 1) was chosen for sampling, located in the inner part of the estuary. Surface sediments (top layer from 0 to 5 cm depth) with their initial water content were collected using a plastic bucket, previously washed with 10% HNO3. In the Huelva estuary, due to seawater depth, professional divers helped on the sampling procedure. The water depths at each sampled site were 1.9 m for MZ and 4.9 m for LEV. In the Río San Pedro estuary, sediment was collected directly using the plastic bucket since at low tide there is no water overlying the sediment. Sediments were transported directly to the laboratory and sieved through a 2 mm plastic mesh before being stored at 4  C, in darkness until their use (Riba et al., 2004). Elutriates were prepared no longer than one month after sediments collection. The water used in the experiments was obtained from the Río San Pedro estuary at the same site as the RSP sediment (Fig. 1). It was collected at a depth of about 1 m from the surface at the mouth of the Río San Pedro during high tide (salinity 30  1), transported to the laboratory (Marine and Environmental Science Faculty, University of Cádiz), filtered using a high-power external filter and kept in a 400 L tank until use. This water was used for preparing elutriates and as the control sample in the experiments. Here, it is referred as “clean seawater” and its natural pH is 8.0. 2.2. Experimental set-up Non-pressurized laboratory-scale chambers were set up to perform experiments using different pH treatments. The pH in each chamber was adjusted and maintained using an Aqua Medic AT control system. Electrodes for pH measurement were placed in the chambers. The pH values in each chamber were adjusted by a solenoid valve that opens when it is detected that the pH has increased by 0.01 units or more; then, CO2 gas bubbles are injected through a hose placed in each chamber until the desired pH value is reached. The AT control system is connected to a computer by which it is possible to modify the pH. Two different sizes of test chamber were used: one is a 2-L glass beaker, to prepare the elutriates, and the other is a 500 mL glass Erlenmeyer flasks, used for the toxicity tests. 2.3. Elutriate preparation Sediment elutriate was obtained by mixing sediment and filtered seawater (ratio 1:4 v/v) inside 2-L glass beakers. The mix was maintained under agitation (60 rpm) for 30 min, followed by a settling down process for at least 12 h (Beiras, 2002; DelValls et al., 1998). Afterwards, the supernatant fraction was carefully extracted, filtered (0.22 mm) and placed in 500 mL Erlenmeyers for the microalgae toxicity tests. The pH of each elutriate was adjusted before and after mixing, by means of CO2 injection. The pH values chosen for this study were 6.0, 7.0 and 8.0.

Table 1 Concentration of total dissolved metals (mg/L) in sediment elutriate. All metals were below detection limit of the equipment in clean seawater samples. pH

[Fe]

CO2 experiments RSP elutriate 8 8.56 7 8.40 6 9.99 MZ elutriate 8 6.35 7 255 6 1221 LEV elutriate 8 49.9 7 11.1 6 22.1 HCl experiments RSP elutriate 8 <3.00a 7 <3.00a 6 6.48 MZ elutriate 8 <3.00a 7 9.23 6 11.9 LEV elutriate 8 6.69 7 <3.00a 6 31.6 a

[Co]

[Cu]

0.97 0.91 1.74

<2.00a <2.00a <2.00a

2.35 2.65 6.97

<2.00a <2.00a <2.00a

3.59 3.68 5.34

<0.50a 1.00 0.52

[Zn]

[As]

[Pb]

3.30 4.08 3.77

<0.30a <0.30a 0.39

16.5 12.1 35.1

10.5 12.8 8.83

<0.30a <0.30a <0.30a

2.32 3.56 5.25

40.3 48.2 80.8

8.83 6.61 5.53

<0.30a <0.30a 6.07

<2.00a <2.00a <2.00a

2.96 15.6 13.0

<2.00a <2.00a <2.00a

1.36 1.38 14.1

2.53 3.39 2.75

21.4 22.0 15.1

23.2 28.4 36.5

10.6 9.33 8.75

1.42 1.03 2.40

3.35 3.40 6.01

22.2 17.5 8.61

25.0 25.8 40.9

6.36 5.16 5.08

0.61 0.97 1.39

5.63 6.03 9.50

Values are referred to the detection limit of each metal on the equipment.

2.4. Culture conditions Phaeodactylum tricornutum was selected because this specie has been proposed as standard organism for seawater toxicity tests (Moreno-Garrido et al., 2000) and because of its easy cultivation, it is one of the most used algal specie in marine bioassays (Horvati c and Persic, 2007). Phaeodactylum tricornutum was obtained from the Culture Collection of Marine Microalgae of the Instituto de Ciencias Marinas de Andalucía (ICMAN-CSIC). The diatom was pre-cultured in 200 mL filtered sterilized natural seawater that was sampled in the Río San Pedro estuary and enriched with f/2 medium (Guillard and Ryther, 1962) in previously autoclaved glass Erlenmeyer flasks (500 mL). Batch cultures were used throughout. The Erlenmeyer flasks were fitted with transpirable tops and maintained at 22  C under continuous illumination at 120 mmol photon m2 s1. The cultures were shaken gently once a day, to ensure aeration and to reduce cells settlement on the bottom of the flasks. The environmental conditions during the study were in accordance with international standards for such tests (ISO, 1995). 2.5. Toxicity tests Algae were taken from the pre-cultures on exponential growth phase and inoculated in glass Erlenmeyer flasks containing 200 mL of sediment elutriates or clean seawater, which was used as control or non-contaminated sample. The toxicity tests (96 h) were performed in triplicates of batch mode cultures and initial cell density was established at 104 cells/mL. The elutriates and the clean seawater were enriched with SiO2 (50 mg/L), NO 3 (150 mg/L), and PO3 4 (10 mg/L). This seawater enrichment was chosen as culture medium for the toxicity experiments because EDTA, present in f/2 medium, decreases the toxicity of metals due to its chelating properties. The pH of elutriates and clean seawater were maintained for 96 h at the same values used for the elutriate preparation, pH 8 (0.1), pH 7 (0.1) and pH 6 (0.1) by means of CO2 injection

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or HCl addition. The CO2 used in the experiments was industrial grade CO2 (99% purity) and the concentration used in each pH treatment is presented in Table 2. For HCl experiments, pH buffer was used (102 M HEPES or MES) to maintain the pH constant. However, every 24 h pH values were checked and, if necessary, more HCl was added. Two solutions of analytical grade HCl (37% Panreac Ref. 141020.1612.4) were prepared for adjusting pHs: 1 M and 0.1 M. After 96 h, samples were collected to determine cell density that was quantified in a Neubauer chamber in an optical microscope, as well as for chlorophylla analysis that was determined according to U.S. EPA (1997), while its concentration was computed from the equation of Jeffrey and Humphrey (1975).

Division (ICP/AAS) of the “Servicios Centrales de Ciencia y Tecnología” of the University of Cádiz. The equilibrium chemicalspeciation/mass transfer model PHREEQC (Parkhurst and Appelo, 1999) was used to determine metal speciation in the obtained elutriates. Thermodynamic database WATEQ4F (Ball and Nordstrom, 1991) was applied in all the calculations. Input data included measured values of metals, temperature, salinity, pH (NBS scale), alkalinity and estimated concentrations of major elements  þ 2þ 2þ þ  (SO2 4 , Cl , K , Na , Mg , Br , F ), which were calculated by the salinity values and considering the stoichiometry of the major solute components; they were based on the assumption that the relative composition of seawater is constant (Millero and Sohn, 1992).

2.6. Elutriate analysis

2.7. Statistical analysis

Elutriate conductivity was measured by a Conductivity 3000 H meter (accuracy 1% of measured value). Total alkalinity (TA) was determined by automatic titration (Metrohm 794) using a combined glass electrode (Metrohm, ref. 6.0210.100) calibrated on a NBS scale. The initial values of TA and pH were used to determine the carbonate system speciation using the CO2SYS program (Pierrot et al., 2006) with dissociation constant from Mehrbach et al. (1973) and refit by Dickson and Millero (1987) and KHSO4 according to Dickson (1990). Seawater and elutriate were filtered (0.45 mm) and acidified to pH < 2 with ultrapure grade HNO3 (2%) for metals analysis and concentration of Fe, Co, Cu, Zn, Pb, Cd, Ni, Cr and As were determined by inductively coupled plasma-mass spectrometry (ICP-MS) (Thermo Elemental Series-X) performed at the Spectroscopy

Differences between acidity induced by HCl versus CO2 were analysed performing t-test using STATISTICA 6.0 software. The results of the elutriate analysis as well as toxicity values were used to conduct a factor analysis, 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. 3. Results 3.1. Chemical composition of the elutriates The initial metal content in elutriates is shown in Table 1, whereas Table 2 includes data for seawater and elutriate carbonate

Table 2 Carbonate system speciation in seawater pH treatments and sediment elutriate pH treatments for CO2 and HCl experiments. pH (NBS scale) CO2 experiments Control (seawater) 7.97 7.05 6.11 RSP elutriate 7.81 6.93 6.23 MZ elutriate 7.84 6.90 6.19 LEV elutriate 7.77 6.87 6.15 HCl experiments Control (seawater) 8.07 7.17 6.26 RSP elutriate 8.13 7.18 6.28 MZ elutriate 8.12 7.17 6.27 LEV elutriate 8.07 7.17 6.27

TA (mmol/kg)

TIC (mmol/kg)

HCO 3 (mmol/kg)

CO2 3 (mmol/kg)

CO2 (mmol/kg)

pCO2 (matm)

Ucalcite

Uaragonite

2754 2544 2481

2565 2690 4493

2400 2412 2506

144 26.0 1.60

21.0 252 1985

341 4037 29,434

3.47 0.39 0.00

2.18 0.25 0.00

2497 3096 4713

2387 3426 7583

2263 3057 4704

94.0 16.0 4.00

30.0 354 2875

487 5648 45,836

2.25 0.38 0.10

1.33 0.24 0.06

3029 3423 5086

2894 3834 8534

2742 3389 5077

119 14.0 4.00

34.0 431 3453

537 6887 55,168

2.85 0.34 0.10

1.79 0.21 0.06

2692 2792 3753

2669 3258 6498

2549 2771 3747

59.0 8.00 2.70

61.0 479 2749

977 7658 43,922

1.40 0.20 0.06

0.88 0.13 0.04

11,322 6515 6088

10,566 6889 9548

9754 6405 6076

745 50.0 6.00

67.0 434 3466

1078 6927 55,372

18.0 1.20 0.13

11.0 0.75 0.08

11,636 6000 6106

10,741 6315 9418

9816 5892 6093

866 49.0 6.00

59.0 374 3319

939 5978 53,027

21.0 1.17 0.14

13.0 0.74 0.09

12,717 6000 6167

11,767 6326 9590

10,773 5894 6154

928 48.0 6.00

66.0 384 3430

1054 6131 54,809

22.0 1.15 0.14

14.0 0.72 0.09

11,498 6052 6153

10,730 6398 9568

9906 5949 6140

756 46.0 6.00

68.0 403 3422

1094 6433 54,684

18.0 1.11 0.14

11.0 0.70 0.09

The carbon parameters were calculated based on initial values for pH, temperature (20  C), salinity (30), and total alkalinity.

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Fig. 2. Phaeodactylum tricornutum final densities. Data represent the mean value for each elutriate and pH treatment tested, including standard deviations, n ¼ 3.

system speciation. The chemical composition of elutriates differed markedly not only between different sampling sites but also between different pH values, and among samples acidified by either CO2 bubbling or HCl addition. Data presented in Table 1 show that sediments from all sampling sites enriched seawater with Fe, As, Co, Pb and Zn, while Cu was only present at the elutriates from the Huelva estuary. All metals were below detection limit of the equipment in clean seawater, as well as the metals Cd, Cr, and Ni in all samples, therefore these data are not included in Table 1. In general, trace metal concentrations were higher in elutriates of sediments from Huelva estuary, with exception of Pb. Sediment acidification increased the release of the metals Fe, Co, Pb and Zn from the sediment to elutriates, nevertheless, this response varied depending on metal, sampling site and also between methods of acidification (CO2 or HCl). The mobility of Fe was higher in elutriates acidified by means of CO2 bubbling than by those acidified with HCl addition, especially in elutriates from MZ site. Furthermore, in the mentioned elutriate, acidification had great influence in metal release, since at the lowest pH tested (pH 6) concentrations of Fe was about three orders of magnitude higher than at the control pH (pH 8). Changes in pH values did not have a clear influence in the metal As either considering the sampling station or the method of acidification. Although Co concentrations were low in all samples, it increased with acidification, especially in samples from elutriates obtained by CO2 bubbling. For Cu, concentrations were higher in elutriates acidified through HCl addition when compared with CO2 bubbling, however, the control pH (pH 8) presented higher Cu concentrations than lower pH (pH 6). Lead concentration was under the detection limit in all samples from CO2 experiments, except for LEV at pH 6, while for HCl samples, acidification resulted in the release of this metal in RSP site. Amounts of Zn differed between samples from elutriate obtained

by CO2 bubbling and HCl addition, nevertheless, acidification have increased the release of this metal in both methods used, as well as in all studied sites, with higher concentrations in elutriate from LEV. 3.2. Microalgae response Fig. 2 shows the algal density after 96 h exposure for the different elutriate and pH tested. The result for chlorophyll-a followed the same pattern as algal density and therefore only the first

Table 3 Microalgae densities and percentage of responses. pH Growth (106 cells/ mL) Control CO2 8 1.75 7 2.00 6 0.53 RSP CO2 8 1.29 7 2.22 6 0.02 MAZ CO2 8 1.79 7 1.42 6 0.00 LEV CO2 8 2.79 7 2.28 6 0.01

Response (%)

Growth (106 cells/ mL)

Response (%)

15 80

HCl 1.85 1.42 0.01

24 99

26 27 99

HCl 1.15 1.40 0.11

38 24 94

3 18 100

HCl 0.75 1.28 0.05

59 31 97

60 31 99

HCl 0.85 1.25 0.07

54 32 96

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parameter was chosen to be illustrated. t-Test analysis showed significant differences of microalgae responses between HCl and CO2 experiments in all elutriates tested as well as in the natural seawater (control). For pH 8, these differences were observed only in the contaminated sites, while for pH 7 and 6 it was also observed in the control and in the elutriate from RSP. Cell number and percentage of response after exposure in the different pH tested are presented in Table 3. The percentage of response was calculated through the equation [(Ns  Nc)/Nc]  100, where Ns, is the cell number in the sample and Nc is the cell number in the control (Mucha et al., 2003). Samples from clean seawater in its original pH (pH 8.0) were considered as control, two different values were used, one for CO2 and another for HCl. Negative response values mean inhibition of algal growth while positive values means stimulation of algal growth. In HCl experiments all response values were negative meaning inhibition of algal growth in all elutriates and pH tested, furthermore, in the elutriates, values are more negative at pH 8 than at pH 7 indicating higher inhibition at this pH, while for pH 6 in all elutriates and seawater, inhibition was almost 100%. On the other hand, in CO2 experiments, microalgae growth was stimulated in all samples at pH 7, except for elutriates from MZ site, while in pH 8 elutriate from LEV site resulted in 60% stimulation of algal growth. For pH 6 microalgae was greatly inhibited in all samples, especially in elutriates. 3.3. Multivariate analysis A factor analysis was conducted to correlate acidification, contamination and the toxicity variables, and to determine the influence of these correlation on each one of the sampling stations. Three new factors were obtained from the original values (Table 4). These factors explained together 77% of the variance. The predominant factor (Factor 1) described 38% of the variance and associated significantly microalgae responses (final biomass and chlorophyll-a) with pH, through negative values. Factor 2 accounted for 24% of the variance and showed microalgae responses (final biomass and chlorophyll-a) associated with metals Co, Cu, Zn and As and pH. The weight of each factor in each sampling station is shown in Table 5. 4. Discussion It is well known that the factor controlling biological effects of metals is their chemical speciation, rather than its total concentration (Allen, 1993; Lombardi et al., 2002; Nogueira et al., 2005). Usually, bioavailability and toxicity of metals to aquatic biota are related to their free forms (Nogueira et al., 2005; Lombardi and Maldonado, 2011). Changes in environmental factors, such as acidification due to increases in CO2 concentrations can change the Table 4 Values of the factors representing initial variables after multivariate analysis of the results obtained. Only the values >0.30. are shown.

pH Fe Co Cu Zn As Pb Microalgae density Chlorophyll-a

F1 (38%)

F2 (24%)

F3 (14%)

0.616 0.474 0.789 0.396 0.709 0.483 0.403 0.764 0.755

0.531

e 0.777 e 0.642 e e 0.437 e e

e 0.527 0.406 0.443 0.723 0.446 0.565 0.482

141

Table 5 Relevance of each factor in the different samples. Only the values >0.30. are shown. Site CO2 experiments Control (seawater) pH 8 pH 7 pH 6 RSP elutriate pH 8 pH 7 pH 6 MZ elutriate pH 8 pH 7 pH 6 LEV elutriate pH 8 pH 7 pH 6 HCl Control (seawater) pH 8 pH 7 pH 6 RSP elutriate pH 8 pH 7 pH 6 MZ elutriate pH 8 pH 7 pH 6 LEV elutriate pH 8 pH 7 pH 6

F1

F2

F3

0.63 0.66 0.95

1.25 0.87 0.68

1.50 1.08

0.57 0.57 0.69

0.53 e e

0.85 0.80 1.29

0.5 0.96 1.04

1.28 2.32

2.21 1.92 0.55

1.12 1.53 2.70

0.37 0.15 1.12

0.38 e 1.27

0.61 0.76 1.09

1.23 0.95 0.81

0.36 0.65

0.65 0.13 0.92

0.67 0.69 e

0.98 e 1.11

0.17 0.39 0.19

1.71 1.55 1.24

1.07 e 0.98

0.59 0.70 1.56

0.35 e 0.47

e 0.90 e e

e

e

speciation and therefore the behaviour and fate of metals (Millero et al., 2009). However, other factors can interfere with the toxicity of metals to marine organisms, for instance, protons can compete with the toxic metal for the binding sites associated with the biota so that toxicity can decrease as pH falls (Tipping et al., 2003). For microalgae, both processes can be observed in literature. In one hand, metal toxicity increases at lower pH due to the prevalence of free-metal ions (Rai et al., 1993; Starodub et al., 1987) and on the other hand, competition between metal ion and Hþ for cell surface has increased toxicity at higher pH (De Schamphelaere et al., 2003; Franklin et al., 2000; Macfie et al., 1994; Wilde et al., 2006). In order to clarify which processes were responsible for microalgae response on each of our study sites, a PCA analysis was performed. Furthermore, geochemical modelling was applied to have information about the species of metals present in each elutriate, as well as to have insights on their interactions with and effects on metal toxicity to microalgae. The results of analysis of prevalence shown in Table 5, indicates by the weight of factor 1, which correlates microalgae response with pH, that microalgae response to acidity was significant in the elutriates from all sampling stations at pH 6. In the samples acidified by means of CO2 injection, this response was more significant in the elutriates, while in samples acidified by means of HCl addition, this factor was also high in the control station. This analysis suggests that an important factor controlling inhibition of algal growth at pH 6 was acidity, however, the presence of metals was also influencing microalgae growth. These are supported by higher cell density and chlorophyll-a in clean seawater than in elutriates, mainly in CO2 experiments (Table 3). The effect of CO2 enrichment on P. tricornutum has been studied before, however since the focus of the study was scenarios of ocean

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acidification due to the uptake of CO2, minimum pH value was 7.8, and in that case, enhancement of growth by 5% was observed (Wu et al., 2010). Other diatom, such as Attheya sp. (King et al., 2011) was also stimulated by CO2 addition while growth of Thalassiosira pseudonana was not influenced (Yang and Gao, 2012) but, again, in all cases minimum pH was around 7.8. In the present study, at pH 7, growth stimulation was also observed and detected through biomass increase at 96 h; in clean seawater, P. tricornutum density was 15% higher at pH 7 than at pH 8. Nevertheless this is only true for CO2 acidification experiments, while for HCl, at pH 7 algae was inhibited by 24%. These differences in microalgae responses may be because membrane permeability for Hþ ions and/or competition for binding sites in its surface are different from the gaseous CO2, and therefore can exert different physiological influences in marine organisms (Heisler, 1986; Morris et al., 1989). The weight of factor 2, which correlates pH values and metal ions with microalgae responses, was significant only for elutriates from the contaminated sites. Furthermore, the weight of this factor was stronger at pH 6, especially in samples from CO2 experiment. This analysis corroborates the hypothesis that inhibition of algal growth and chlorophyll-a synthesis in the contaminated elutriates is owed by two main factors: acidity and toxicity of metals. 4.1. Toxic metals Within the metals included in factor 2 (As, Co, Cu, Zn) Cu is probably the most studied regarding toxicity to microalgae. For P. tricornutum, low EC50 values have been reported in literature, such as 8 mg/l (Levy et al., 2008). This concentration is similar to those found in this research in elutriates from LEV and MZ sites acidified by means of HCl addition, whereas elutriates from CO2 bubbling had lower Cu values. Higher Cu concentration in HCl samples could explain the inhibition of algae growth in LEV (pH 7 and pH 8) and MZ (pH 8) sites if compared with same samples in CO2 bubbling, even though for MZ site factor 2 was not correlated at pH 8. Furthermore, samples from pH 8 suffered more inhibition than those at pH 7, which is not surprising since, as previously mentioned, metal toxicity for microalgae have been reported to be higher at higher pH values (De Schamphelaere et al., 2003; Franklin et al., 2000; Macfie et al., 1994; Wilde et al., 2006). Moreover, for Cu, it has been observed that decreases in pH have increased the amount of Cu2þ even though, toxicity decreased (Wilde et al., 2006). In this study, geochemical modelling demonstrated increases in the concentration of free Cu2þ ions with acidification in both HCl and CO2 elutriates. While at pH 8 Cu2þ was not present in elutriates, at pH 6 it represented around 10% of total Cu. It should be considered here that about 99.8 a 99.9% of the copper in seawater can be associated with natural DOM (Sunda and Hanson, 1987) and because this DOM was most likely present in our experiments, we can consider that at least part of such Cu2þ may have gone to DOM complexes, known to be less toxic to phytoplankton when compared with free Cu2þ ions (Lombardi et al., 2002, 2005). Regarding the other metals included in factor 2, As has a complex chemistry that involves chemical, biochemical and geochemical reactions that have important implications on its toxicity to marine organisms and their consumers, including humans (Neff, 1997). Geochemical modelling applied suggests acidification did not influence the speciation of this metal. Arsenic was present as As (V), mainly as HAsO2 4 , in all sampling sites and pH tested. There is a controversy in literature on which species of arsenic (arsenite or arsenate) is more toxic to microalgae (Karadjova et al., 2008). According to Foster et al. (2008), arsenate is the actual specie considered to be taken up by microalgae from seawater via the phosphate transport systems located in cell membranes and afterwards converted to As (III). The range of concentration of As

producing toxic effects in microalgae in literature is usually much higher than those found in our elutriates (Ismail et al., 2002; Satoh et al., 2005). However few works have reported similar or even lower values of arsenate producing toxicity to green algae and diatoms (Karadjova et al., 2008; Sanders, 1979). In addition, it has been shown that decreases in pH can increase the toxicity of As (V) as observed in experiments with Stichococcus bacillaris (PawlikSkowronska et al., 2004), therefore, even though concentrations of As in our elutriates did not increase due to acidification, its toxicity could have been increased, and As could be one of the metals responsible for the toxic effects to microalgae at pH 6. While total concentration of Co was slightly influenced by acidification, its speciation was not affected by pH changes and nearly 90% of Co was present as its free form, Co2þ. Concentration of this metal was very low in all samples and therefore Co is probably not responsible for toxicity. Within the metals included in factor 2, Zn suffered more influence of acidification. Concentration of this metal in pH 6 in elutriates from contaminated sites was twice higher than at pH 7 and 8. Values for Zn toxicity to microalgae in literature vary from mg/l (Johnson et al., 2007; Schamphelaere et al., 2005) to mg/l (Monteiro et al., 2011) range. Toxic values are usually higher than those present in our elutriates, however, inhibitory responses have been observed in the range of values reported here (Schamphelaere et al., 2005; Wilde et al., 2006). Regarding Zn speciation, acidification increased the concentration of Zn2þ, while at pH 8 free form of Zn represented less than 40% of total zinc, at pH 7 and 6 it represented almost 50%. Since at pH 6 concentrations of total Zn were higher than in other pHs, and free Zn also increased we suggest that Zn can also have contributed for microalgae growth inhibition at the lowest pH elutriates. 4.2. Iron Iron can be a limiting factor for photosynthesis, thus, having direct effects in microalgae community, and to be used by phytoplankton, it must be dissolved. It is known that decreases in pH, increases the solubility of Fe (III) (Liu and Millero, 2002), however, there is a reliance that Fe (II) is the actual specie taken up by phytoplankton, but the latter species is generally not considered as an abundant source of bioavailable iron due to its short residence time in oxygenated waters. The reduced form of iron usually suffers rapid reoxidation to Fe (III), however, these oxidation reactions are strongly dependent upon pH (Millero et al., 2009). For example, it is expected that a pH decrease of 0.5 units in seawater results in a 10fold increase in the half-life of Fe (II) (Breitbarth et al., 2010). The output of geochemical model applied in this work suggests that the concentration of reduced form of iron increased when pH decreased. While at pH 8, almost all of Fe was found as Fe (III), at the lowest pH tested (pH 6), concentration of Fe (II) represented more than 70% of total Fe in all sampling sites. For pH 7 differences were found between elutriates acidified by HCl and CO2 methods. In samples from CO2 experiment, Fe (II) represented almost 50% of total iron at this pH value, while in HCl samples, reduced form of iron represented less than 10%. The presence of Fe might have stimulated microalgae growth in elutriates from CO2 experiments, where, at pH 8 from LEV site, the highest Fe concentration at this pH coincided with higher cell density. Under our experimental conditions, higher total Fe and higher Fe (II) were detected in the CO2 systems when compared with the HCl systems. As reported in Table 2, the higher Uarag detected in HCl systems (UaragHCl) can have contributed to Fe/CaCO3 precipitates formation, so reducing Fe in such systems, whereas in the CO2 systems ðUaragCO2 Þ this was possibly less intense due to the lower UaragCO2 . This rationale applies specially for pH 7.0 and 8.0

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(exception MZ pH6), where the difference between HCl and CO2 systems was higher and also were the differences in diatom growth; considering pH 6.0, UaragHCl w UaragCO2 . At the same time that we detected lower Fe, higher bicarbonate was present in the HCl systems, nevertheless lower biomass was obtained, possibly due to lower Fe and higher Cu. Still, based on the results presented in Table 2, we can conclude in a general view that HCl induced acidity was more effective in unbalancing the seawater carbonate system and toxic metal ions solubilization than was the acidity induced by CO2 increment. This could have lead to lower diatom growth in the HCl in comparison with the CO2 systems. So, considering an environmental point of view, CO2 acidification can end up resulting in increased eutrophication of marine ecosystem, and consequently to all the undesired side reactions, such as substitution of phytoplankton species with increase in cyanobacteria and flagellates, together with the final reduction of biodiversity. 5. Conclusion Strategies to mitigate climate change, such as the CO2 capture and storage into marine geological formations is currently receiving much attention. However, the risks associated with this activity have to be carefully investigated in order to construct robust database and perform correct impact assessment of such activity. The experiments performed in this investigation showed that sediment acidification resulted in the release of metals to elutriates. Iron and zinc were the metals most influenced by this process and their concentration increased greatly with pH decrease. Geochemical modelling showed that acidification also influenced the speciation of the metals Fe and Cu, enhancing the concentration of their most bioavailable form in the elutriates. Differences in metal release between experiments performed with CO2 bubbling and HCl addition were observed, and these differences also reflected distinct microalgae responses between both acidification methodologies. Toxicity of metals to microalgae at several pH were greater when pH was manipulated using HCl, resulting in negative response values in all pH treatments and sites. In CO2 experiments, higher algal density in clean seawater than in elutriates, at pH 6, suggested inhibition of algae was not attributed only to acidity but also to the presence of metals, while multivariate analysis of our data together with literature database lead us to propose that a combination of factors, such as unbalance of seawater carbonate system and the solubilization of metals such as Zn were related to the toxicity. On the other hand, Cu provoked toxicity at higher pH in HCl samples, and this metal was associated with differences in algal growth between HCl and CO2 samples. Moreover, the increase of Fe due to CO2 acidification influenced microalgae response as well, but this time producing stimulatory effects. Geochemical modelling showed that CO2 acidification did not influence calcite and aragonite formation as did the HCL acidification, so Fe is likely to remain in solution in CO2, but precipitate in HCl systems. This can have important ecological consequences such as eutrophication of nearby waters, leading to phytoplankton species substitution and affect biodiversity. Our results showed that CO2 release from CCS activities is likely to result in release of toxic metals, as well as iron to seawater and also to interfere in marine organisms such as microalgae. Furthermore, combined effects of acidity and the presence of metals increased these interferences. Data on the influence of ocean acidification in marine organisms, including microalgae, are growing in literature, however, most works are focused on acidification regarding the direct uptake of CO2, and since leakage from CO2 seabed storage sites would create faster and stronger acidification than by those induced by atmospheric CO2, more studies using different organisms and contaminants in the context of CCS

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are needed in order to quantify the potential effects of this activities. Acknowledgements The first author thanks CAPES/MEC-Brazil (BEX-0-501/09-3) for the doctoral scholarship. Additionally, the work was partially funded by the Spanish ministry of education projects (CTM201128437-C02-02/TECNO and CTM2012-36476-C02-01/tecno). ATL thanks CNPq e Brazil for a research scholarship (302837/2012-4) and ARR thanks Junta de Andalucía (Regional Government) for grant reference RNM-3924. References Allen, H.E., 1993. The significance of trace metal speciation for water, sediment and soil quality criteria and standards. Sci. Total Environ. 134, 23e45. Ardelan, M.V., Steinnes, E., Lierhagen, S., Linde, S.O., 2009. Effects of experimental CO2 leakage on solubility and transport of seven trace metals in seawater and sediment. Sci. Total Environ. 407, 6255e6266. Ardelan, M.V., Steinnes, A., 2010. Changes in the mobility and solubility of the redox sensitive metals Fe, Mn, and Co at the seawater-sediment interface following CO2 seepage. Biogeosciences 7, 1e15. Ball, J.W., Nordstrom, D.K., 1991. User’s Manual for WATEQ4F, with Revised Thermodynamic Data Base and Test Cases for Calculating Speciation of Major, Trace, and Redox Elements in Natural Waters. US Geol. Surv. Open-File Rep, pp. 91e 183. Basallote, M.D., Rodríguez-Romero, A., Blasco, J., DelValls, A., Riba, I., 2012. Lethal effects on different marine organisms, associated with sedimenteseawater acidification deriving from CO2 leakage. Environ. Sci. Pollut. Res. 19, 2550e2560. Beiras, R., 2002. Comparison of methods to obtain a liquid phase in marine sediment toxicity bioassays with Paracentrotus lividus sea urchin embryos. Arch. Environ. Contam. Toxicol. 42, 23e28. Blackford, J., Jones, N., Proctor, R., Holt, R., Widdicombe, S., Lowe, D., Rees, A., 2009. An initial assessment of the potential environmental impact of CO2 escape from marine carbon capture and storage systems. Proc. Inst. Mech. Eng. A: J. Power and Energy 223, 269e280. Blasco, J., 2005. Estudio de la Recuperación del Caño de Sancti Petri despues de la Desaparición de los Vertidos de Aguas Residuales de La Ciudad de San Fernando. Final Report Project. Consejeria Medio Ambiente. Junta de Andalucía. Ministerio de Educación y Ciencia, Spain. Breitbarth, E., Achterberg, E.P., Ardelan, M.V., Baker, A.R., Bucciarelli, E., Chever, F., Croot, P.L., Duggen, S., Gledhill, M., Hassellov, M., Hassler, C., Hoffmann, L.J., Hunter, K.A., Hutchins, D.A., Ingri, J., Jickells, T., Lohan, M.C., Nielsdóttir, M.C., Sarthou, G., Schoemann, V., Trapp, J.M., Turner, D.R., Ye, Y., 2010. Iron biogeochemistry across marine systems e progress from the past decade. Biogeosciences 7, 1075e1097. De Orte, M.R., Sarmiento, A.M., Basallote, M.D., Rodriguez-Romero, A., Riba, I., DelValls, A., 2013. Effects on the mobility of metals from acidification caused by possible CO2 leakage from sub-seabed geological formations. Sci. Total Environ. (in press). De Schamphelaere, K.A.C., Vasconcelos, F.M., Heijerick, D.G., Tack, F.M.G., Delbeke, K., Allen, H.E., Janssen, C.R., 2003. Development and field validation of a predictive copper toxicity model for the green alga Pseudokirchneriella subcapitata. Environ. Toxicol. Chem. 22, 2454e2465. DelValls, T.A., Blasco, J., Sarasquete, M.C., Forja, J.M., Gómez-Parra, A., 1998. Evaluation of heavy metal sediment toxicity in littoral ecosystems using juveniles of the fish Sparus aurata. Ecotoxicol. Environ. Saf. 41, 157e167. Dickson, A.G., Millero, F.J., 1987. A comparison of the equilibrium constants for the dissociation of carbonic acid in seawater media. Deep-sea Res. 34, 1733e 1743. Dickson, A.G., 1990. Standard potential of the reaction: AgCl(s) þ 1/2 H2(g) ¼ Ag(s) þ HCl(aq), and the standard acidity constant of the ion HSO4 in synthetic seawater from 273.15 to 318.15 K. J. Chem. Thermodyn. 22, 113e127. Franklin, N.M., Stauber, J.L., Markich, S.J., Lim, R.P., 2000. pH-dependent toxicity of copper and uranium to a tropical freshwater alga (Chlorella sp.). Aquat. Toxicol. 48, 275e289. Franqueira, F., Orosa, M., Torres, E., Herrero, C., Cid, A., 2000. Potential use of flow cytometry in toxicity studies with microalgae. Sci. Total Environ. 247, 119e126. Foster, S., Thomson, D., Maher, W., 2008. Uptake and metabolism of arsenate by axenic cultures of the microalgae Dunaliella tertiolecta and Phaeodactylum tricornutum. Mar. Chem. 108, 172e183. Guillard, R.R.L., Ryther, J.H., 1962. Studies on marine planktonic diatoms, I. Cyclotella nana Hustedt and Detonula confervaceae (Cleve) Gran. Can. J. Microbiol. 8, 229e 239. Hale, R., Calosi, P., McNeill, L., Mieszkowska, N., Widdicombe, S., 2011. Predicted levels of future ocean acidification and temperature rise could alter community structure and biodiversity in marine benthic communities. Oikos 120, 661e674. Heisler, N., 1986. Acid-base Regulation in Animals. Elsevier, Amsterdam.

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