BIOJEC-06696; No of Pages 7 Bioelectrochemistry xxx (2013) xxx–xxx
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Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes Marie Libert a,⁎, Marta Kerber Schütz a, Loïc Esnault b, Damien Féron c, Olivier Bildstein a a b c
CEA, DEN, DTN/SMTM/LMTE, F-13108 St Paul lez Durance, France ECOGEOSAFE, Technopole environnement Arbois-Méditerranée, Avenue Louis Philibert, F-13545 Aix-en-Provence Cedex, France CEA, DEN, DANS/DPC, F-91191 Gif-sur-Yvette, France
a r t i c l e
i n f o
Article history: Received 13 February 2013 Received in revised form 30 September 2013 Accepted 2 October 2013 Available online xxxx Keywords: Geological disposal Biocorrosion Sulphate-reducing bacteria Iron-reducing bacteria Dihydrogen
a b s t r a c t This study emphasizes different experimental approaches and provides perspectives to apprehend biocorrosion phenomena in the specific disposal environment by investigating microbial activity with regard to the modification of corrosion rate, which in turn can have an impact on the safety of radioactive waste geological disposal. It is found that iron-reducing bacteria are able to use corrosion products such as iron oxides and “dihydrogen” as new energy sources, especially in the disposal environment which contains low amounts of organic matter. Moreover, in the case of sulphate-reducing bacteria, the results show that mixed aerobic and anaerobic conditions are the most hazardous for stainless steel materials, a situation which is likely to occur in the early stage of a geological disposal. Finally, an integrated methodological approach is applied to validate the understanding of the complex processes and to design experiments aiming at the acquisition of kinetic data used in long term predictive modelling of biocorrosion processes. © 2013 Elsevier B.V. All rights reserved.
1. Introduction The generally accepted strategy to manage high level and long lived radioactive waste (HLLW) is the disposal in deep stable geological formations (i.e. clay or granitic host-rocks). The containment concept is based on multi-barriers with various materials, such as vitrified and bituminized waste, metal containers (carbon steel, stainless steel), bentonite and concrete confined in the geological medium [1]. In France, the National Radioactive Waste Management Agency (Andra) considers an argillaceous site, studied by means of the Underground Research Laboratory (URL) in Bure (Meuse, France), for the commissioning of the geological disposal in 2025. The assessment of the long term behaviour of the radioactive waste and disposal materials is required in order to demonstrate the safety of this strategy and to assure the future geological disposal. The major vector of alteration is the water coming from the site, which will be in physical contact with the waste after corrosion of the steel components (lining of disposal cells and containers). Therefore, the study of the phenomena linked to corrosion in deep geological conditions aims at assessing the containment capacity of the metallic materials. The observation of an important microbiological diversity in these conditions [2–5] leads us also to consider the impact of these microorganisms in terms of corrosion. In spite of the extreme ⁎ Corresponding author. Tel.: +33 4 42253144; fax: +33 4 42256272. E-mail address: [email protected] (M. Libert).
physicochemical conditions prevailing during the early stage of radioactive waste disposal (redox conditions, high pressure and temperature, radiation), microorganisms still show a strong capacity for survival and adaptation. The introduction of exogenic materials in the geological medium is also likely to significantly modify the geochemical conditions and to provide chemical compounds that will enable microbial activities. Moreover, a significant amount of dihydrogen (H2) will be produced due to the corrosion of metal containers and/or radiolysis of water. H2 will constitute an energetic substrate for anaerobic bacterial communities, especially in this environment containing only small amounts of organic matter [6], and will strongly stimulate reducing reactions involving Fe(III) and sulphates [7–10]. Fe(III) mineral phases such as magnetite (Fe3O4) (a corrosion product) will provide electron acceptors, further supporting the development of iron-reducing bacteria (IRB) [11,12]. Note that this study concerns only metallic materials, even though biological impact (dissolution, biodegradation, redox reactions on iron mineral phases) has been demonstrated on other materials used in geological repositories, such as concrete, bitumen and claystone [10,13]. The main objective of this paper is to give an overview of the different approaches applied to the long term prediction of the (bio) corrosion processes and to discuss experimental results from recent studies concerning the impact of sulphate-reducing and iron-reducing microbial activities on the mechanisms responsible for corrosion, such as alteration of corrosion products (e.g. magnetite), and production/ consumption of gas (e.g. H2).
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Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
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2. Overview 2.1. Long term prediction In the framework of HLLW disposal, the relevant timescale is several thousands to several hundreds of thousands of years. Therefore, the semi-empirical modelling often used to evaluate the lifetime of materials at the scale of industrial facilities is no longer sufficient. Tackling the challenge of understanding and predicting the corrosion phenomena and kinetics, over such a long duration, calls for an integrative approach including: - laboratory or in situ experiments in order to understand the phenomena, to identify the physicochemical mechanisms, and to determine the key parameters controlling the corrosion rate (over short periods of time); - integrated laboratory or in situ experiments conducted in representative conditions of the repository with complete coupling of the effects of the geological medium, metallic material, waste form, etc; - archaeological analogues in order to check the long term prediction in terms of values for the kinetics parameters and the validity of the mechanisms determined previously with short term experiments; - mechanistic modelling calibrated at the scale of the laboratory experiments (based on the physicochemical processes), and extrapolated to long term prediction of the evolution of large scale system. The study of the biocorrosion phenomena in deep geological disposal is thus aimed at the long term prediction approach, which consists, as a first stage, in the identification of the physicochemical mechanisms due to the strong interactions between the microorganisms, disposal materials and geological medium, as described previously [14]. Scheme 1 summarizes this “3M approach” which is detailed in the following sections. 2.2. Interactions between microorganisms, disposal materials and geological medium Even though activity and development of microorganisms in compacted clay medium are still a matter of debate (the radius of pores may be too small) [15], it will certainly be possible in the technological gaps and in the connected fractured zone (Fig. 1).
In the first hundreds of years of the repository lifetime, the physicochemical conditions may not be favourable for microbiological development, especially during the thermal phase where irradiation, high temperature (up to 90 °C) and low water saturation will prevail. However, bacterial activity will resume after this period when the conditions are back to temperature between 20–70 °C, radioactivity b60 Sv/h, and water saturation N0.97 [1,16]. 2.2.1. The geological media According to Gaucher et al. (2004), the physicochemical conditions and the concentration of nutrients seem to be favourable for bacterial development in the case of the Callovo-Oxfordian claystone formation in the Bure site [17]. The different stages of the repository lifetime (exploitation, closure with a hydraulic oxidized transient phase and then reduced conditions [1]) will also strongly influence the geochemical conditions in the geological media. Esnault et al. (2013) determined the concentration of potential energetic and nutrient substrates (C, N, P, S) during the lifetime of the repository for the Bure site [18]. Their results indicate the presence of key elements in significant amounts, such as sulphates (10−2 mol/L). The elements for which the concentration most likely limits the bacterial activity are phosphates and nitrates. The claystone in the Bure URL contains only low amounts of organic matter. In contrast, there are several mineral phases which may also provide nutrients for bacterial activity (mainly Fe(II) from pyrite, Fe(III) from clay minerals and carbonate minerals). Fig. 1 summarizes the location of potential bacterial activity in the near field and the processes providing nutrients and energetic substrates. 2.2.2. The disposal materials The introduction of exogenic materials such as steel and glass in the geological repository will significantly modify the inventory of the energetic and nutrients substrates as well as the redox conditions by generating H2 [1] and Fe(III) bearing corrosion products (i.e. magnetite) [19]. These elements will control the selection and the development of specific bacterial activities in this particular environment. 2.2.3. The microorganisms The existence of microorganisms in deep geological conditions has been largely described [2–5]. The nutrients provided by the host-rock and the disposal materials as well as the physicochemical conditions will control the type of metabolisms which are likely to develop and
Scheme 1. Interactions between microorganisms, disposal materials and geological medium.
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
M. Libert et al. / Bioelectrochemistry xxx (2013) xxx–xxx
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Fig. 1. Schematic view of the near field of a HLLW disposal cell indicating the location of potential microbial activity and the flux of nutrients and energetic substrates (not to the scale).
their activity, independently from the origin of microorganisms which will be present in the disposal environment (exogenous or endogenous bacteria). The significant amounts of H2 expected into the repository will support the development of hydrogenotrophic bacteria, which use H2 as electron donor for energy source [10]. This type of metabolism is frequently found in environments poor in biodegradable organic matter [20]. Therefore, these hydrogenotrophic bacteria will couple H2 oxidation to the dissimilatory reduction of elements such as sulphates (sulphatereducing bacteria, SRB) or Fe(III) (IRB) from (hydr)oxide minerals formed during the corrosion process [11] and from Fe(III) bearing clay minerals [13,18]. The consumption of H2 by bacteria seems rather a favourable effect because it will decrease the build-up of H2 gas expected into the repository (it can damage the barrier properties of the geological formation) and the eventual risks of H2 embrittlement of the metallic materials. In contrast, it has to be noticed that the accumulation of H2 may also decrease the free corrosion potential and the kinetics of iron dissolution [21,22]. 3. Materials and methods The impact of microbial activities on corrosion can be evaluated by means of electrochemical and geochemical methods. 3.1. SRB experiments — Electrochemical methods The materials and experimental conditions are extensively described in [23] for the evaluation of breakdown potentials (crevice corrosion potentials) of stainless steels with and without SRB in seawaters. The experiments were performed with several types of seawaters: (i) natural seawater taken from the harbour of Genova (Italy); and (ii) sterilised seawater made with natural seawater treated for two days with about 1–2 μM of chlorine; then chlorine was eliminated with a suitable sodium sulphite addition and the solution was aerated for sterilised aerated seawater. SRB samples were taken from a natural source (natural anaerobic marine mud, natural inoculum) and then inoculated with de-aerated sterilised seawater. The austenitic stainless steel was a 316L type (17.0% of chromium, 11.5% of nickel, 2.2% of molybdenum, 0.06% of nitrogen and less than 0.03% of carbon). Working electrodes were made with stainless coupons (50 × 50 × 1.5 mm) equipped with crevice formers made of rubber rings (Ø 15 mm, thickness 2 mm).
Standard electrochemical systems with 3 electrodes were used for these experiments (working electrodes, counter electrode made of platinum and a standard calomel electrode (SCE) as reference). Several working electrodes made of SS coupons equipped with crevices (up to a maximum of 60 coupons) were concurrently exposed to the test solutions (about 100 L) and connected to a data logger and to a potentiostat. After a week, when all parameters (pH, sulphide concentration, free corrosion potentials) were stabilized, anodic polarizations from −500 to +500 mV/SCE, and return, with a scanning rate of 1.23 V/day (1 mV/70 s), were concurrently applied to evaluate breakdown potentials.
3.2. IRB experiments — Geochemical methods The experiments with IRB species were followed using geochemical indicators, which consists in monitoring the concentration of dissolved iron and H2 as a function of time. This approach was recently applied to a facultative anaerobic bacterium Shewanella oneidensis strain MR-1 (IRB model organism) [11,24]. The experiments with and without bacteria were performed in sterile anaerobic batch reactors containing 140 mL of a synthetic medium representative of the Callovo-Oxfordian claystone porewater from the Bure site. The chemical composition is the following: 9 mM (NH4)2SO4, 0.33 mM K2HPO4, 0.33 mM KH2PO4, 2 mM NaHCO3, 1 mM MgSO4.7H2O, 0.4 mM CaCl2.2H2O, 45 μM H3BO3, 10 μM NaCl, 4 μM FeSO4·7H2O, 5 μM CoSO4.7H2O, 5 μM NiSO4·6H2O, 4μM Na2MoO4·2H2O, 11 μM Na2SeO4, 1.3 μM MnSO4·H2O, 1 μM ZnSO4·7H2O, 0.2 μM CuSO4·5H2O, 0.115 mM arginine, 0.2 mM glutamate, 0.2 mM serine, and a vitamin solution (0.08 mM nicotinic acid, 0.01 mM thiamine, 0.4 μM biotine). The pH was adjusted to 7 with NaOH (0.1 N). The medium was sterilised by autoclaving (120°C for 20min), except for the thermolabile components (e.g. amino acids) which were filter-sterilised (0.22 μm) and added to the autoclaved medium. Cultures of S. oneidensis strain MR-1 (ATCC 700550™) were obtained aerobically at the beginning of the stationary growth phase in a Luria Bertani Broth (LB) medium (5 g/L NaCl, 10 g/L tryptone, 5 g/L yeast extract) after 24 h at 30 °C under sterile conditions. Bacterial cells were harvested from the LB medium by centrifugation (4000 rpm for 20 min), washed once with the sterile synthetic medium described above and then inoculated in the biotic reactors (initial concentration 108 cells mL−1 counted by epifluorescence method with LIVE/DEAD® BacLight™ kit).
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
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1 g of metallic iron powder (Merck, Ø 10 μm) or magnetite powder synthesized according to [25] was added into the reactors. Iron powder was sterilised in an irradiator at the CEA Cadarache Centre (cobalt radioactive source (60Co) delivering a total dose of 35 kGy to kill live bacteria) and magnetite powder by UV irradiation for 30min. Anaerobic conditions were obtained by bubbling a gas mixture: 90% N2, 10% CO2 for iron powder experiments and 60% H2, 30% N2, 10% CO2 for magnetite experiments. Abiotic experiments (without bacteria) were conducted in the same conditions. All reactors were then incubated at 30 °C. The duration of experiment was about 1000 h for iron powder experiments and about 600h for magnetite powder experiments. Three replicates for each abiotic or biotic experiment were performed. The graphs of the results show the average of these three replicates. The aqueous Fe concentration was analyzed by Inductively Coupled Plasma Optical Emission Spectrometry (ICP-OES, Varian, VISTA-MPX) after 2% (v/v) HNO3 acidification. The analysis of H2 in the headspace was carried out by Micro Gas Chromatography (Varian, CP-4900) using a thermal conductivity detector with N2 as carrier gas. The total H2 concentration was calculated as the sum of the concentrations in the gaseous and the aqueous phases (determined using Henry's law). The corrosion products formed in the experiments were characterized by X-Ray diffraction (XRD). XRD scans were recorded in the angular range of 2°–40° (2θ) (counting time of 2 s per 0.02° (2θ) step size) with a Bruker D8 diffractometer (Cu radiation at 40 kV and 40 mA) equipped with a Göbel mirror and a Sol-X detector. 4. Results and discussion 4.1. Laboratory experiments 4.1.1. Sulphate-reducing bacteria (SRB) The literature describes that in the absence of oxygen, the corrosion rate is low if we consider nonacidic aqueous solutions. The rate is then limited by the cathodic reaction (Eq. (1)): −
2H2 O þ 2e →H2 þ 2OH
−
ð1Þ
accompanied by the anodic reaction (Eq. (2)): 2þ
Fe→Fe
−
þ 2e
ð2Þ
In our experiments the follow-up of free corrosion potential of the stainless steel samples highlights that in presence of SRB their electrochemical behaviour is primarily related to the sulphides content and to the pH, rather than to the presence of SRB. For example, there is no difference between the free corrosion potential in the presence/ absence of Desulfovibrio vulgaris or Desulfovibrio gigas, in the case where the chemical composition (sulphides content and pH) are identical in both conditions [23]. With regard to the potentials of pitting corrosion, additional experiments were performed in order to verify if the cumulative distribution of the breakdown potentials for crevice corrosion changes in presence of SRB. To support this study, the culture medium for SRB was substituted with de-aerated seawater in which suitable doses of Na2S were added in such a way that the same total sulphide concentration and the same pH of the SRB culture were obtained. Fig. 2 shows the comparison of the cumulative distribution of the breakdown potentials (crevice corrosion potentials expressed in volts versus SCE), obtained in a SRB culture, as specified in §3.1, and in a Na2S deaerated seawater solution, respectively; total sulphide concentration was close to 12.5 mM and pH close to 7 in both solutions [23]. It can be seen that no appreciable differences are observed between the two cumulative distributions. It should be pointed out that the values of these breakdown potentials are very low in the presence of sulphides (12.5 mM in the example of Fig. 2) or SRB species. In conclusion, sulphides added as Na2S to sterile de-aerated seawater can simulate the effect of biologically generated sulphides, provided that
Fig. 2. Cumulative distribution of breakdown potentials (f, expressed as the statistical frequency) as function of the crevice corrosion potentials (CCP, volts (V) versus saturated calomel electrode (SCE)) for stainless steel 316L in SRB culture (♦) and in deaerated seawater added with an equivalent amount of sulphides (12.5 mM of S2− added as Na2S) (▪) and with same pH = 7 at room temperature (scanning rate: 175 mV/day).
pH and total sulphide concentration are equal in both solutions. The effect of SRB on the stainless steel samples is thus primarily related to the chemical modifications (sulphides content and pH) which they impose in the surrounding medium. Another issue which is often raised in terms of corrosion is the coexistence of aerobic and anaerobic conditions [23]. Fig. 3 shows how the data can be used to predict the role of the presence/absence of aerobic/anaerobic biofilms in promoting the onset of crevice corrosion. The effect of SRB on the stainless steel samples decreases dramatically the breakdown potentials: in natural aerated seawater without sulphide, the crevice corrosion potentials are around +0.3 V/SCE (Fig. 3b), while the crevice corrosion potentials are around −0.2 V/SCE in the presence of sulphide (12.5 mM) or SRB (Fig. 3c). Fig. 3a summarises the behaviour of the stainless steel samples exposed to sterile seawater. Taking into account that the maximum free corrosion potential in this solution is lower than +100 mV/SCE, the position of the cumulative distribution of the breakdown potentials suggests that the spontaneous onset of crevice corrosion is a very improbable phenomenon in these conditions. Fig. 3b and c describes two possible behaviour of the stainless steel samples: in the first case (Fig. 3b) exposed to natural seawater in the presence of an aerobic young biofilm alone (few days or few weeks biofilm); and, in the second case (Fig. 3c), in the presence of an old biofilm (few months of few years old biofilm) with both aerobic and anaerobic zones on the shielded area. In the first case, the comparison between the maximum free corrosion potential values and the position of the cumulative distribution of the breakdown potentials for crevice corrosion on the shielded areas (in absence of sulphides) indicates that the spontaneous onset of crevice corrosion must be considered as a relatively frequent phenomenon. Obviously, the probability of spontaneous crevice corrosion onset in the presence of aerobic biofilm decreases if a more alloyed stainless steel type is tested. Finally, in the second case, aerobic and anaerobic zones in the biofilm act in synergy making absolutely sure that the spontaneous onset of crevice corrosion will occur. Mixed redox conditions are the most susceptible ones for these passivable materials: their free corrosion potential will be increased by the action of the aerobic bacteria in the biofilm, whereas the presence of SRB in the anaerobic niches of the biofilm will result in a strong decrease of the localized corrosion resistance of the alloy. In other terms, the cathodic reaction is increased by the aerobic biofilm, whereas the anodic curve is affected by the presence of SRB. The increase of the cathodic reaction rate leads to an increase in the free corrosion potential, whereas the presence of sulphides involves a reduction in the breakdown potential in the zones where SRB are present.
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
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4.1.2. Iron-reducing bacteria (IRB) The role of IRB on biocorrosion phenomena has seldom been studied to date. In the first studies conducted in the context of geological disposal [11–13,24], the impact of IRB activity was evaluated by geochemical methods, as previously mentioned. According to the results from [11], IRB species can have an impact on the corrosion products formed during the corrosion process, as shown in Fig. 4. The results showed that the production of H2 is accompanied by the release of iron in solution (in the first 100 h). After 100 h, the production of H2 decreases because of the precipitation of iron oxides (magnetite), which forms a passive layer. The addition of bacteria at approximately 400 h is followed by an increase of dissolved iron due to the “bioreduction” of Fe(III) from magnetite. At the same time, a resumption of the H2 production is observed which is attributed to a reactivation of the metallic corrosion process. Recent studies have investigated the impact of IRB on the corrosion of low carbon steel combining geochemical and electrochemical approaches [24,27]. These studies confirm the enhancement of the corrosion rate (by a factor of 2–3) due to the bacterial activity during short term batch experiments. According to [27], it was observed that this enhancement of corrosion concerns a generalized phenomenon, as no localized corrosion was observed in the presence of IRB. In addition, recent experiments have shown the consumption of H2 by S. oneidensis coupled to Fe(III) reduction from synthetic magnetite (Fig. 5). Previous results highlight that the microbial activity may alter iron passive layers made of Fe(III). As a result of Fe(III) reduction, a precipitation of FeCO3 can be observed [11]. This paragenesis is also presented in experiments using carbon steel in contact with an argillite core sample [19]. 4.2. Integrated experiments Integrated laboratory experiments (column reactors with steel coupons in contact with an argillite core sample) described previously in the literature have been performed with SRB [28,29] or mixed bacterial populations (IRB and SRB) [30] in representative conditions of geological disposal, under pressure and ventilated medium/closed systems. The results confirm the survival and growth of microorganisms in disposal conditions, but their activity could be limited by the porosity of clay media. Despite the fact that clay porosity is dominated by pore sizes of about 10 nm, the presence of heterogeneities or cracks (due to excavation) and the remaining technological gaps represent favourable space volumes for bacterial development. Quantification of the biocorrosion rate in such experiments needs in situ specific integrated experiments. 4.3. Archaeological analogues Fig. 3. Cumulative distribution of breakdown potentials (f, expressed as the statistical frequency) as function of the crevice corrosion potentials (CCP, volts (V) versus saturated calomel electrode (SCE) for stainless steel 316L under sterile seawater(Fig. 3a), in presence of aerobic young biofilm in aerated natural seawater (Fig. 3b) and in presence of an old biofilm with both aerobic and anaerobic zones (Fig. 3c) at room temperature.
It is noteworthy to stress that the environment of the geological repository will successively be in oxidized conditions (during the disposal exploitation phase), then in anoxic and reduced conditions after the disposal cells closure. The transition between these two phases, where the metallic materials can simultaneously be in oxidized conditions in some locations and in reduced conditions in others, is certainly a matter of concern since localized corrosion could be triggered by either differential oxygen supply, or by the simultaneous presence of aerobic and anaerobic bacteria, as observed in thick sea water biofilms [14,26].
Archaeological analogues found under natural conditions with the presence of SRB confirm that these bacteria are not very aggressive for steel, in case of strictly anaerobic conditions. For example, mean corrosion rates as low as 0.03 μm/year were determined for throws and swords being more than 1700 years old and preserved in marshes [31], which corresponds to a passive behaviour of these materials in terms of corrosion rate. 4.4. Mechanistic and integrated modelling Modelling can be used to calculate the chemical composition of the groundwater (e.g. using an equilibrium assumption for minerals phases, according to the scenarios considered) in order to determine the availability of essential nutrients for microbial development. Integrated simulations on iron–clay interactions show that most of the biological redox processes will be localized close to the radioactive waste containers, where nutrients are immobilized in minerals (in
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
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Fig. 4. Influence of iron-reducing bacteria on metal corrosion: bioreduction of corrosion products. Fe (●,▪) and H2 (▲,♦) concentrations (mM) as function of time (hours). Abiotic conditions (♦,▪) and biotic conditions (▲,●). Aqueous corrosion (1), magnetite precipitation (2), Fe(III) bioreduction from magnetite (3) and reactivation of corrosion process (4).
particular, as phosphates) and where a high amount of H2 and magnetite is produced [18]. Again, the availability of these substrates can especially favour the development of IRB species. Nevertheless, global bacterial activity modelling is complex and requires many parameters to better predict the long term evolution of the disposal system.
effect on the evolution of the repository. Further investigations in this direction involve the implementing of a global bacterial activity model representing the main biological reactions in predictive reactive transport models.
5. Conclusions
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To predict biocorrosion in deep geological disposal conditions, a global approach with specific and integrated experiments is necessary to support a representative and predictive modelling which evaluates the lifetime of the disposal materials at the scale of industrial facilities. In this study, it is demonstrated that IRB are able to use corrosion products such as Fe(II,III) iron oxides (magnetite) and dihydrogen as new energy sources in this specific environment. The results also show that mixed aerobic and anaerobic redox conditions tend to favour biocorrosion in the presence of SRB, which constitutes a probable situation in geological disposal systems. It has been demonstrated that IRB and SRB are implicated in direct or indirect biocorrosion phenomena in the experimental timeframe, but it remains to determine if the bacterial activity is likely to have a long term
Fig. 5. H2 consumption by hydrogenotrophic iron-reducing bacteria under Fe(III) reduction from magnetite. H2 (▲,▪) concentrations (mM) as function of time (days). Abiotic conditions (▪) and biotic conditions (▲). Fe3O4/solution ratio (g) = 1/140.
References
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
M. Libert et al. / Bioelectrochemistry xxx (2013) xxx–xxx [17] E. Gaucher, C. Robelin, J.M. Matray, G. Negral, Y. Gros, J.F. Heitz, A. Vinsot, H. Rebours, A. Cassagnabere, A. Bouchet, ANDRA underground research laboratory: interpretation of the mineralogical and geochemical data acquired in the Callovian–Oxfordian formation by investigative drilling, Phys. Chem. Earth 29 (2004) 55–77. [18] L. Esnault, M. Libert, O. Bildstein, Assessment of microbiological development in nuclear waste geological disposal: a geochemical modelling approach, Procedia Earth Planet. Sci. 7 (2013) 244–247. [19] M.L. Schlegel, C. Bataillon, K. Benhamida, C. Blanc, D. Menut, J.L. Lacour, Metal corrosion and argillite transformation at the water-saturated, high-temperature iron-clay interface: a microscopic-scale study, Appl. Geochem. 23 (2008) 2619–2633. [20] K.H. Nealson, F. Inagaki, K. Takai, Hydrogen-driven subsurface lithoautotrophic microbial ecosystems (SLiMEs): do they exist and why should we care? Trends Microbiol. 13 (2005) 405–410. [21] R. Javaherdashti, A brief review of general patterns of MIC of carbon steel and biodegradation of concrete, IUFS J. Biol. 68 (2009) 65–73. [22] S. Kakooei, M.C. Ismail, B. Ariwahjoedi, Mechanisms of microbiologically influenced corrosion: a review, World Appl. Sci. J. 17 (2012) 524–531. [23] U. Kivisakk, B. Espelid, D. Feron, Methodology of Crevice Corrosion Testing for Stainless Steels in Natural and Treated Seawaters, European Federation of Corrosion (EFC) Series, Maney Publishing, Leeds UK, 2010. [24] M.K. Schütz, R. Moreira, O. Bildstein, J.E. Lartigue, M.L. Schlegel, B. Tribollet, V. Vivier, M. Libert, Combined geochemical and electrochemical methodology to quantify corrosion of carbon steel by bacterial activity, Bioelectrochem. (2013), http: //dx.doi.org/10.1016/j.bioelechem.2013.07.003(in this issue). [25] U. Schwertmann, R.M. Cornell, Iron Oxides in the Laboratory: Preparation and Characterization, VCH, Weinheim, New York, 1991. (137 pp.). [26] I.B. Beech, J. Sunner, Biocorrosion: towards understanding interactions between biofilms and metals, Curr. Opin. Biotechnol. 15 (2004) 81–186. [27] R. Moreira, M.K. Schütz, M. Libert, B. Tribollet, V. Vivier, Local electrochemical measurements applied to biocorrosion, Bioelectrochem. (2013), http: //dx.doi.org/10.1016/j.bioelechem.2013.10.003(in this issue). [28] V. Madina, I. Azkarate, M. Insausti, V. Madina, I. Azkarate, M. Insausti, Corrosion of several components of the in situ test performed in a deep geological granite disposal site, in: prediction of long term corrosion behavior in nuclear waste systems, Science and Technology Series, ANDRA Publisher, France, 2005. 61–67. [29] H. El Hajj, A. Abdelouas, B. Grambow, C. Martin, M. Dion, Microbial corrosion of P235GH steel under geological conditions, Phys. Chem. Earth 35 (2010) 248–253. [30] C. Chautard, J.E. Lartigue, M. Libert, F. Marsal, L. De Windt, An integrated experiment coupling iron/argillite interactions with bacterial activity, Procedia Chem. 7 (2012) 641–646. [31] H. Matthiesen, L.R. Hilbert, D. Gregory, B. Sarensen, Long term corrosion of iron at the waterlogged side Nydam in Denmark: studies of environment, archeological artefacts and modern analogues, in: Prediction of long term corrosion behavior in nuclear waste systems, Science and Technology Series, ANDRA Publisher, France, 2005. 114–127. Marie Libert received her PhD in Biochemistry (1986) from the University of Compiegne-France. She is a senior research engineer at the CEA since 1998. Her previous affiliation includes a doctoral fellowship at the same university. Her research focuses on different fields such as the effect of microorganisms on long term behaviour of materials used in nuclear repository, biocorrosion, impact of radioactive emission on biota, anaerobic microbiology. She was and she is a European expert for several international research programs.
7 Marta Kerber Schütz received a BS degree in Industrial Chemistry from the Catholic University of Rio Grande do Sul, Porto Alegre, Brazil in 2007; completed her MSc in Materials Science and Engineering in March 2010 at the same University. Now she is a PhD student at the French Commission for Atomic Energy and Alternative Energies (CEA). She is working on biocorrosion phenomena implicated on the radioactive waste geological disposal. The overall objective of her study is to better understand the impact of bacterial activities on the corrosion products (specially hydrogen and iron (hydr)oxides) and on the rate of anoxic corrosion.
Loïc Esnault earned a PhD in Geosciences, specialized on biogeochemistry, from Nancy University in 2010. He has obtained a multidisciplinary competence during his scholarship with a Bachelor's degree in Microbiology from the University of Paris 6 “Pierre et Marie Curie” in 2005, a Master's degree in Environmental Engineering and Materials Science from the School of Bridge and Road (ENPC) in 2007. His research activities now focus on the integrity of material and their behaviour in environmental conditions.
Damien Féron received his PhD degree in Chemical Engineering in 1979. Working since more than 30 years in the corrosion field, he is a Research Director at the CEA's Nuclear Energy Directorate and Professor at the National Institute of Nuclear Science and Engineering (INSTN). He has been elected Chairman of the Science and Technology Advisory Committee (STAC) of the European Federation of Corrosion.
Olivier Bildstein received his PhD in Geochemistry (1998) from the University of Strasbourg. He is a senior research engineer at the CEA since 2000, after a postdoctoral stay at LLNL (California, USA). His previous affiliations include a doctoral fellowship at the French Institute for Petroleum (1994–1998) and a position as a hydrogeologist at the French Geological Survey (BRGM) (1993). His research focuses on numerical modelling of reactive, multicomponent, and multiphase flow and transport, with application to long term evolution of materials in radioactive waste disposal, subsurface pollution/remediation, and potential bacteria mediated processes.
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
Contents lists available at ScienceDirect
Bioelectrochemistry journal homepage: www.elsevier.com/locate/bioelechem
Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes Marie Libert a,⁎, Marta Kerber Schütz a, Loïc Esnault b, Damien Féron c, Olivier Bildstein a a b c
CEA, DEN, DTN/SMTM/LMTE, F-13108 St Paul lez Durance, France ECOGEOSAFE, Technopole environnement Arbois-Méditerranée, Avenue Louis Philibert, F-13545 Aix-en-Provence Cedex, France CEA, DEN, DANS/DPC, F-91191 Gif-sur-Yvette, France
a r t i c l e
i n f o
Article history: Received 13 February 2013 Received in revised form 30 September 2013 Accepted 2 October 2013 Available online xxxx Keywords: Geological disposal Biocorrosion Sulphate-reducing bacteria Iron-reducing bacteria Dihydrogen
a b s t r a c t This study emphasizes different experimental approaches and provides perspectives to apprehend biocorrosion phenomena in the specific disposal environment by investigating microbial activity with regard to the modification of corrosion rate, which in turn can have an impact on the safety of radioactive waste geological disposal. It is found that iron-reducing bacteria are able to use corrosion products such as iron oxides and “dihydrogen” as new energy sources, especially in the disposal environment which contains low amounts of organic matter. Moreover, in the case of sulphate-reducing bacteria, the results show that mixed aerobic and anaerobic conditions are the most hazardous for stainless steel materials, a situation which is likely to occur in the early stage of a geological disposal. Finally, an integrated methodological approach is applied to validate the understanding of the complex processes and to design experiments aiming at the acquisition of kinetic data used in long term predictive modelling of biocorrosion processes. © 2013 Elsevier B.V. All rights reserved.
1. Introduction The generally accepted strategy to manage high level and long lived radioactive waste (HLLW) is the disposal in deep stable geological formations (i.e. clay or granitic host-rocks). The containment concept is based on multi-barriers with various materials, such as vitrified and bituminized waste, metal containers (carbon steel, stainless steel), bentonite and concrete confined in the geological medium [1]. In France, the National Radioactive Waste Management Agency (Andra) considers an argillaceous site, studied by means of the Underground Research Laboratory (URL) in Bure (Meuse, France), for the commissioning of the geological disposal in 2025. The assessment of the long term behaviour of the radioactive waste and disposal materials is required in order to demonstrate the safety of this strategy and to assure the future geological disposal. The major vector of alteration is the water coming from the site, which will be in physical contact with the waste after corrosion of the steel components (lining of disposal cells and containers). Therefore, the study of the phenomena linked to corrosion in deep geological conditions aims at assessing the containment capacity of the metallic materials. The observation of an important microbiological diversity in these conditions [2–5] leads us also to consider the impact of these microorganisms in terms of corrosion. In spite of the extreme ⁎ Corresponding author. Tel.: +33 4 42253144; fax: +33 4 42256272. E-mail address: [email protected] (M. Libert).
physicochemical conditions prevailing during the early stage of radioactive waste disposal (redox conditions, high pressure and temperature, radiation), microorganisms still show a strong capacity for survival and adaptation. The introduction of exogenic materials in the geological medium is also likely to significantly modify the geochemical conditions and to provide chemical compounds that will enable microbial activities. Moreover, a significant amount of dihydrogen (H2) will be produced due to the corrosion of metal containers and/or radiolysis of water. H2 will constitute an energetic substrate for anaerobic bacterial communities, especially in this environment containing only small amounts of organic matter [6], and will strongly stimulate reducing reactions involving Fe(III) and sulphates [7–10]. Fe(III) mineral phases such as magnetite (Fe3O4) (a corrosion product) will provide electron acceptors, further supporting the development of iron-reducing bacteria (IRB) [11,12]. Note that this study concerns only metallic materials, even though biological impact (dissolution, biodegradation, redox reactions on iron mineral phases) has been demonstrated on other materials used in geological repositories, such as concrete, bitumen and claystone [10,13]. The main objective of this paper is to give an overview of the different approaches applied to the long term prediction of the (bio) corrosion processes and to discuss experimental results from recent studies concerning the impact of sulphate-reducing and iron-reducing microbial activities on the mechanisms responsible for corrosion, such as alteration of corrosion products (e.g. magnetite), and production/ consumption of gas (e.g. H2).
1567-5394/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
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2. Overview 2.1. Long term prediction In the framework of HLLW disposal, the relevant timescale is several thousands to several hundreds of thousands of years. Therefore, the semi-empirical modelling often used to evaluate the lifetime of materials at the scale of industrial facilities is no longer sufficient. Tackling the challenge of understanding and predicting the corrosion phenomena and kinetics, over such a long duration, calls for an integrative approach including: - laboratory or in situ experiments in order to understand the phenomena, to identify the physicochemical mechanisms, and to determine the key parameters controlling the corrosion rate (over short periods of time); - integrated laboratory or in situ experiments conducted in representative conditions of the repository with complete coupling of the effects of the geological medium, metallic material, waste form, etc; - archaeological analogues in order to check the long term prediction in terms of values for the kinetics parameters and the validity of the mechanisms determined previously with short term experiments; - mechanistic modelling calibrated at the scale of the laboratory experiments (based on the physicochemical processes), and extrapolated to long term prediction of the evolution of large scale system. The study of the biocorrosion phenomena in deep geological disposal is thus aimed at the long term prediction approach, which consists, as a first stage, in the identification of the physicochemical mechanisms due to the strong interactions between the microorganisms, disposal materials and geological medium, as described previously [14]. Scheme 1 summarizes this “3M approach” which is detailed in the following sections. 2.2. Interactions between microorganisms, disposal materials and geological medium Even though activity and development of microorganisms in compacted clay medium are still a matter of debate (the radius of pores may be too small) [15], it will certainly be possible in the technological gaps and in the connected fractured zone (Fig. 1).
In the first hundreds of years of the repository lifetime, the physicochemical conditions may not be favourable for microbiological development, especially during the thermal phase where irradiation, high temperature (up to 90 °C) and low water saturation will prevail. However, bacterial activity will resume after this period when the conditions are back to temperature between 20–70 °C, radioactivity b60 Sv/h, and water saturation N0.97 [1,16]. 2.2.1. The geological media According to Gaucher et al. (2004), the physicochemical conditions and the concentration of nutrients seem to be favourable for bacterial development in the case of the Callovo-Oxfordian claystone formation in the Bure site [17]. The different stages of the repository lifetime (exploitation, closure with a hydraulic oxidized transient phase and then reduced conditions [1]) will also strongly influence the geochemical conditions in the geological media. Esnault et al. (2013) determined the concentration of potential energetic and nutrient substrates (C, N, P, S) during the lifetime of the repository for the Bure site [18]. Their results indicate the presence of key elements in significant amounts, such as sulphates (10−2 mol/L). The elements for which the concentration most likely limits the bacterial activity are phosphates and nitrates. The claystone in the Bure URL contains only low amounts of organic matter. In contrast, there are several mineral phases which may also provide nutrients for bacterial activity (mainly Fe(II) from pyrite, Fe(III) from clay minerals and carbonate minerals). Fig. 1 summarizes the location of potential bacterial activity in the near field and the processes providing nutrients and energetic substrates. 2.2.2. The disposal materials The introduction of exogenic materials such as steel and glass in the geological repository will significantly modify the inventory of the energetic and nutrients substrates as well as the redox conditions by generating H2 [1] and Fe(III) bearing corrosion products (i.e. magnetite) [19]. These elements will control the selection and the development of specific bacterial activities in this particular environment. 2.2.3. The microorganisms The existence of microorganisms in deep geological conditions has been largely described [2–5]. The nutrients provided by the host-rock and the disposal materials as well as the physicochemical conditions will control the type of metabolisms which are likely to develop and
Scheme 1. Interactions between microorganisms, disposal materials and geological medium.
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
M. Libert et al. / Bioelectrochemistry xxx (2013) xxx–xxx
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Fig. 1. Schematic view of the near field of a HLLW disposal cell indicating the location of potential microbial activity and the flux of nutrients and energetic substrates (not to the scale).
their activity, independently from the origin of microorganisms which will be present in the disposal environment (exogenous or endogenous bacteria). The significant amounts of H2 expected into the repository will support the development of hydrogenotrophic bacteria, which use H2 as electron donor for energy source [10]. This type of metabolism is frequently found in environments poor in biodegradable organic matter [20]. Therefore, these hydrogenotrophic bacteria will couple H2 oxidation to the dissimilatory reduction of elements such as sulphates (sulphatereducing bacteria, SRB) or Fe(III) (IRB) from (hydr)oxide minerals formed during the corrosion process [11] and from Fe(III) bearing clay minerals [13,18]. The consumption of H2 by bacteria seems rather a favourable effect because it will decrease the build-up of H2 gas expected into the repository (it can damage the barrier properties of the geological formation) and the eventual risks of H2 embrittlement of the metallic materials. In contrast, it has to be noticed that the accumulation of H2 may also decrease the free corrosion potential and the kinetics of iron dissolution [21,22]. 3. Materials and methods The impact of microbial activities on corrosion can be evaluated by means of electrochemical and geochemical methods. 3.1. SRB experiments — Electrochemical methods The materials and experimental conditions are extensively described in [23] for the evaluation of breakdown potentials (crevice corrosion potentials) of stainless steels with and without SRB in seawaters. The experiments were performed with several types of seawaters: (i) natural seawater taken from the harbour of Genova (Italy); and (ii) sterilised seawater made with natural seawater treated for two days with about 1–2 μM of chlorine; then chlorine was eliminated with a suitable sodium sulphite addition and the solution was aerated for sterilised aerated seawater. SRB samples were taken from a natural source (natural anaerobic marine mud, natural inoculum) and then inoculated with de-aerated sterilised seawater. The austenitic stainless steel was a 316L type (17.0% of chromium, 11.5% of nickel, 2.2% of molybdenum, 0.06% of nitrogen and less than 0.03% of carbon). Working electrodes were made with stainless coupons (50 × 50 × 1.5 mm) equipped with crevice formers made of rubber rings (Ø 15 mm, thickness 2 mm).
Standard electrochemical systems with 3 electrodes were used for these experiments (working electrodes, counter electrode made of platinum and a standard calomel electrode (SCE) as reference). Several working electrodes made of SS coupons equipped with crevices (up to a maximum of 60 coupons) were concurrently exposed to the test solutions (about 100 L) and connected to a data logger and to a potentiostat. After a week, when all parameters (pH, sulphide concentration, free corrosion potentials) were stabilized, anodic polarizations from −500 to +500 mV/SCE, and return, with a scanning rate of 1.23 V/day (1 mV/70 s), were concurrently applied to evaluate breakdown potentials.
3.2. IRB experiments — Geochemical methods The experiments with IRB species were followed using geochemical indicators, which consists in monitoring the concentration of dissolved iron and H2 as a function of time. This approach was recently applied to a facultative anaerobic bacterium Shewanella oneidensis strain MR-1 (IRB model organism) [11,24]. The experiments with and without bacteria were performed in sterile anaerobic batch reactors containing 140 mL of a synthetic medium representative of the Callovo-Oxfordian claystone porewater from the Bure site. The chemical composition is the following: 9 mM (NH4)2SO4, 0.33 mM K2HPO4, 0.33 mM KH2PO4, 2 mM NaHCO3, 1 mM MgSO4.7H2O, 0.4 mM CaCl2.2H2O, 45 μM H3BO3, 10 μM NaCl, 4 μM FeSO4·7H2O, 5 μM CoSO4.7H2O, 5 μM NiSO4·6H2O, 4μM Na2MoO4·2H2O, 11 μM Na2SeO4, 1.3 μM MnSO4·H2O, 1 μM ZnSO4·7H2O, 0.2 μM CuSO4·5H2O, 0.115 mM arginine, 0.2 mM glutamate, 0.2 mM serine, and a vitamin solution (0.08 mM nicotinic acid, 0.01 mM thiamine, 0.4 μM biotine). The pH was adjusted to 7 with NaOH (0.1 N). The medium was sterilised by autoclaving (120°C for 20min), except for the thermolabile components (e.g. amino acids) which were filter-sterilised (0.22 μm) and added to the autoclaved medium. Cultures of S. oneidensis strain MR-1 (ATCC 700550™) were obtained aerobically at the beginning of the stationary growth phase in a Luria Bertani Broth (LB) medium (5 g/L NaCl, 10 g/L tryptone, 5 g/L yeast extract) after 24 h at 30 °C under sterile conditions. Bacterial cells were harvested from the LB medium by centrifugation (4000 rpm for 20 min), washed once with the sterile synthetic medium described above and then inoculated in the biotic reactors (initial concentration 108 cells mL−1 counted by epifluorescence method with LIVE/DEAD® BacLight™ kit).
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
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1 g of metallic iron powder (Merck, Ø 10 μm) or magnetite powder synthesized according to [25] was added into the reactors. Iron powder was sterilised in an irradiator at the CEA Cadarache Centre (cobalt radioactive source (60Co) delivering a total dose of 35 kGy to kill live bacteria) and magnetite powder by UV irradiation for 30min. Anaerobic conditions were obtained by bubbling a gas mixture: 90% N2, 10% CO2 for iron powder experiments and 60% H2, 30% N2, 10% CO2 for magnetite experiments. Abiotic experiments (without bacteria) were conducted in the same conditions. All reactors were then incubated at 30 °C. The duration of experiment was about 1000 h for iron powder experiments and about 600h for magnetite powder experiments. Three replicates for each abiotic or biotic experiment were performed. The graphs of the results show the average of these three replicates. The aqueous Fe concentration was analyzed by Inductively Coupled Plasma Optical Emission Spectrometry (ICP-OES, Varian, VISTA-MPX) after 2% (v/v) HNO3 acidification. The analysis of H2 in the headspace was carried out by Micro Gas Chromatography (Varian, CP-4900) using a thermal conductivity detector with N2 as carrier gas. The total H2 concentration was calculated as the sum of the concentrations in the gaseous and the aqueous phases (determined using Henry's law). The corrosion products formed in the experiments were characterized by X-Ray diffraction (XRD). XRD scans were recorded in the angular range of 2°–40° (2θ) (counting time of 2 s per 0.02° (2θ) step size) with a Bruker D8 diffractometer (Cu radiation at 40 kV and 40 mA) equipped with a Göbel mirror and a Sol-X detector. 4. Results and discussion 4.1. Laboratory experiments 4.1.1. Sulphate-reducing bacteria (SRB) The literature describes that in the absence of oxygen, the corrosion rate is low if we consider nonacidic aqueous solutions. The rate is then limited by the cathodic reaction (Eq. (1)): −
2H2 O þ 2e →H2 þ 2OH
−
ð1Þ
accompanied by the anodic reaction (Eq. (2)): 2þ
Fe→Fe
−
þ 2e
ð2Þ
In our experiments the follow-up of free corrosion potential of the stainless steel samples highlights that in presence of SRB their electrochemical behaviour is primarily related to the sulphides content and to the pH, rather than to the presence of SRB. For example, there is no difference between the free corrosion potential in the presence/ absence of Desulfovibrio vulgaris or Desulfovibrio gigas, in the case where the chemical composition (sulphides content and pH) are identical in both conditions [23]. With regard to the potentials of pitting corrosion, additional experiments were performed in order to verify if the cumulative distribution of the breakdown potentials for crevice corrosion changes in presence of SRB. To support this study, the culture medium for SRB was substituted with de-aerated seawater in which suitable doses of Na2S were added in such a way that the same total sulphide concentration and the same pH of the SRB culture were obtained. Fig. 2 shows the comparison of the cumulative distribution of the breakdown potentials (crevice corrosion potentials expressed in volts versus SCE), obtained in a SRB culture, as specified in §3.1, and in a Na2S deaerated seawater solution, respectively; total sulphide concentration was close to 12.5 mM and pH close to 7 in both solutions [23]. It can be seen that no appreciable differences are observed between the two cumulative distributions. It should be pointed out that the values of these breakdown potentials are very low in the presence of sulphides (12.5 mM in the example of Fig. 2) or SRB species. In conclusion, sulphides added as Na2S to sterile de-aerated seawater can simulate the effect of biologically generated sulphides, provided that
Fig. 2. Cumulative distribution of breakdown potentials (f, expressed as the statistical frequency) as function of the crevice corrosion potentials (CCP, volts (V) versus saturated calomel electrode (SCE)) for stainless steel 316L in SRB culture (♦) and in deaerated seawater added with an equivalent amount of sulphides (12.5 mM of S2− added as Na2S) (▪) and with same pH = 7 at room temperature (scanning rate: 175 mV/day).
pH and total sulphide concentration are equal in both solutions. The effect of SRB on the stainless steel samples is thus primarily related to the chemical modifications (sulphides content and pH) which they impose in the surrounding medium. Another issue which is often raised in terms of corrosion is the coexistence of aerobic and anaerobic conditions [23]. Fig. 3 shows how the data can be used to predict the role of the presence/absence of aerobic/anaerobic biofilms in promoting the onset of crevice corrosion. The effect of SRB on the stainless steel samples decreases dramatically the breakdown potentials: in natural aerated seawater without sulphide, the crevice corrosion potentials are around +0.3 V/SCE (Fig. 3b), while the crevice corrosion potentials are around −0.2 V/SCE in the presence of sulphide (12.5 mM) or SRB (Fig. 3c). Fig. 3a summarises the behaviour of the stainless steel samples exposed to sterile seawater. Taking into account that the maximum free corrosion potential in this solution is lower than +100 mV/SCE, the position of the cumulative distribution of the breakdown potentials suggests that the spontaneous onset of crevice corrosion is a very improbable phenomenon in these conditions. Fig. 3b and c describes two possible behaviour of the stainless steel samples: in the first case (Fig. 3b) exposed to natural seawater in the presence of an aerobic young biofilm alone (few days or few weeks biofilm); and, in the second case (Fig. 3c), in the presence of an old biofilm (few months of few years old biofilm) with both aerobic and anaerobic zones on the shielded area. In the first case, the comparison between the maximum free corrosion potential values and the position of the cumulative distribution of the breakdown potentials for crevice corrosion on the shielded areas (in absence of sulphides) indicates that the spontaneous onset of crevice corrosion must be considered as a relatively frequent phenomenon. Obviously, the probability of spontaneous crevice corrosion onset in the presence of aerobic biofilm decreases if a more alloyed stainless steel type is tested. Finally, in the second case, aerobic and anaerobic zones in the biofilm act in synergy making absolutely sure that the spontaneous onset of crevice corrosion will occur. Mixed redox conditions are the most susceptible ones for these passivable materials: their free corrosion potential will be increased by the action of the aerobic bacteria in the biofilm, whereas the presence of SRB in the anaerobic niches of the biofilm will result in a strong decrease of the localized corrosion resistance of the alloy. In other terms, the cathodic reaction is increased by the aerobic biofilm, whereas the anodic curve is affected by the presence of SRB. The increase of the cathodic reaction rate leads to an increase in the free corrosion potential, whereas the presence of sulphides involves a reduction in the breakdown potential in the zones where SRB are present.
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
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4.1.2. Iron-reducing bacteria (IRB) The role of IRB on biocorrosion phenomena has seldom been studied to date. In the first studies conducted in the context of geological disposal [11–13,24], the impact of IRB activity was evaluated by geochemical methods, as previously mentioned. According to the results from [11], IRB species can have an impact on the corrosion products formed during the corrosion process, as shown in Fig. 4. The results showed that the production of H2 is accompanied by the release of iron in solution (in the first 100 h). After 100 h, the production of H2 decreases because of the precipitation of iron oxides (magnetite), which forms a passive layer. The addition of bacteria at approximately 400 h is followed by an increase of dissolved iron due to the “bioreduction” of Fe(III) from magnetite. At the same time, a resumption of the H2 production is observed which is attributed to a reactivation of the metallic corrosion process. Recent studies have investigated the impact of IRB on the corrosion of low carbon steel combining geochemical and electrochemical approaches [24,27]. These studies confirm the enhancement of the corrosion rate (by a factor of 2–3) due to the bacterial activity during short term batch experiments. According to [27], it was observed that this enhancement of corrosion concerns a generalized phenomenon, as no localized corrosion was observed in the presence of IRB. In addition, recent experiments have shown the consumption of H2 by S. oneidensis coupled to Fe(III) reduction from synthetic magnetite (Fig. 5). Previous results highlight that the microbial activity may alter iron passive layers made of Fe(III). As a result of Fe(III) reduction, a precipitation of FeCO3 can be observed [11]. This paragenesis is also presented in experiments using carbon steel in contact with an argillite core sample [19]. 4.2. Integrated experiments Integrated laboratory experiments (column reactors with steel coupons in contact with an argillite core sample) described previously in the literature have been performed with SRB [28,29] or mixed bacterial populations (IRB and SRB) [30] in representative conditions of geological disposal, under pressure and ventilated medium/closed systems. The results confirm the survival and growth of microorganisms in disposal conditions, but their activity could be limited by the porosity of clay media. Despite the fact that clay porosity is dominated by pore sizes of about 10 nm, the presence of heterogeneities or cracks (due to excavation) and the remaining technological gaps represent favourable space volumes for bacterial development. Quantification of the biocorrosion rate in such experiments needs in situ specific integrated experiments. 4.3. Archaeological analogues Fig. 3. Cumulative distribution of breakdown potentials (f, expressed as the statistical frequency) as function of the crevice corrosion potentials (CCP, volts (V) versus saturated calomel electrode (SCE) for stainless steel 316L under sterile seawater(Fig. 3a), in presence of aerobic young biofilm in aerated natural seawater (Fig. 3b) and in presence of an old biofilm with both aerobic and anaerobic zones (Fig. 3c) at room temperature.
It is noteworthy to stress that the environment of the geological repository will successively be in oxidized conditions (during the disposal exploitation phase), then in anoxic and reduced conditions after the disposal cells closure. The transition between these two phases, where the metallic materials can simultaneously be in oxidized conditions in some locations and in reduced conditions in others, is certainly a matter of concern since localized corrosion could be triggered by either differential oxygen supply, or by the simultaneous presence of aerobic and anaerobic bacteria, as observed in thick sea water biofilms [14,26].
Archaeological analogues found under natural conditions with the presence of SRB confirm that these bacteria are not very aggressive for steel, in case of strictly anaerobic conditions. For example, mean corrosion rates as low as 0.03 μm/year were determined for throws and swords being more than 1700 years old and preserved in marshes [31], which corresponds to a passive behaviour of these materials in terms of corrosion rate. 4.4. Mechanistic and integrated modelling Modelling can be used to calculate the chemical composition of the groundwater (e.g. using an equilibrium assumption for minerals phases, according to the scenarios considered) in order to determine the availability of essential nutrients for microbial development. Integrated simulations on iron–clay interactions show that most of the biological redox processes will be localized close to the radioactive waste containers, where nutrients are immobilized in minerals (in
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
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Fig. 4. Influence of iron-reducing bacteria on metal corrosion: bioreduction of corrosion products. Fe (●,▪) and H2 (▲,♦) concentrations (mM) as function of time (hours). Abiotic conditions (♦,▪) and biotic conditions (▲,●). Aqueous corrosion (1), magnetite precipitation (2), Fe(III) bioreduction from magnetite (3) and reactivation of corrosion process (4).
particular, as phosphates) and where a high amount of H2 and magnetite is produced [18]. Again, the availability of these substrates can especially favour the development of IRB species. Nevertheless, global bacterial activity modelling is complex and requires many parameters to better predict the long term evolution of the disposal system.
effect on the evolution of the repository. Further investigations in this direction involve the implementing of a global bacterial activity model representing the main biological reactions in predictive reactive transport models.
5. Conclusions
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To predict biocorrosion in deep geological disposal conditions, a global approach with specific and integrated experiments is necessary to support a representative and predictive modelling which evaluates the lifetime of the disposal materials at the scale of industrial facilities. In this study, it is demonstrated that IRB are able to use corrosion products such as Fe(II,III) iron oxides (magnetite) and dihydrogen as new energy sources in this specific environment. The results also show that mixed aerobic and anaerobic redox conditions tend to favour biocorrosion in the presence of SRB, which constitutes a probable situation in geological disposal systems. It has been demonstrated that IRB and SRB are implicated in direct or indirect biocorrosion phenomena in the experimental timeframe, but it remains to determine if the bacterial activity is likely to have a long term
Fig. 5. H2 consumption by hydrogenotrophic iron-reducing bacteria under Fe(III) reduction from magnetite. H2 (▲,▪) concentrations (mM) as function of time (days). Abiotic conditions (▪) and biotic conditions (▲). Fe3O4/solution ratio (g) = 1/140.
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7 Marta Kerber Schütz received a BS degree in Industrial Chemistry from the Catholic University of Rio Grande do Sul, Porto Alegre, Brazil in 2007; completed her MSc in Materials Science and Engineering in March 2010 at the same University. Now she is a PhD student at the French Commission for Atomic Energy and Alternative Energies (CEA). She is working on biocorrosion phenomena implicated on the radioactive waste geological disposal. The overall objective of her study is to better understand the impact of bacterial activities on the corrosion products (specially hydrogen and iron (hydr)oxides) and on the rate of anoxic corrosion.
Loïc Esnault earned a PhD in Geosciences, specialized on biogeochemistry, from Nancy University in 2010. He has obtained a multidisciplinary competence during his scholarship with a Bachelor's degree in Microbiology from the University of Paris 6 “Pierre et Marie Curie” in 2005, a Master's degree in Environmental Engineering and Materials Science from the School of Bridge and Road (ENPC) in 2007. His research activities now focus on the integrity of material and their behaviour in environmental conditions.
Damien Féron received his PhD degree in Chemical Engineering in 1979. Working since more than 30 years in the corrosion field, he is a Research Director at the CEA's Nuclear Energy Directorate and Professor at the National Institute of Nuclear Science and Engineering (INSTN). He has been elected Chairman of the Science and Technology Advisory Committee (STAC) of the European Federation of Corrosion.
Olivier Bildstein received his PhD in Geochemistry (1998) from the University of Strasbourg. He is a senior research engineer at the CEA since 2000, after a postdoctoral stay at LLNL (California, USA). His previous affiliations include a doctoral fellowship at the French Institute for Petroleum (1994–1998) and a position as a hydrogeologist at the French Geological Survey (BRGM) (1993). His research focuses on numerical modelling of reactive, multicomponent, and multiphase flow and transport, with application to long term evolution of materials in radioactive waste disposal, subsurface pollution/remediation, and potential bacteria mediated processes.
Please cite this article as: M. Libert, et al., Impact of microbial activity on the radioactive waste disposal: long term prediction of biocorrosion processes, Bioelectrochemistry (2013), http://dx.doi.org/10.1016/j.bioelechem.2013.10.001
