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Coupled Geochemical Impacts of Leaking CO2 and Contaminants from Subsurface Storage Reservoirs on Groundwater Quality Hongbo Shao,*,† Nikolla P. Qafoku, Amanda R. Lawter, Mark E. Bowden, and Christopher F. Brown Pacific Northwest National Laboratory, Richland, Washington 99352, United States S Supporting Information *

ABSTRACT: The leakage of CO2 and the concomitant brine from deep storage reservoirs to overlying groundwater aquifers is considered one of the major potential risks associated with geologic CO2 sequestration (GCS). In this work both batch and column experiments were conducted to determine the fate of trace metals in groundwater in the scenarios of CO2 and metal-contaminated brine leakage. The sediments for this study were from an unconsolidated sand and gravel aquifer in Kansas, containing 0−4 wt % carbonates. Cd (114 μg/L) and As (40 μg/L) were spiked into the reaction system to represent potential contaminants from the reservoir brine. Through this research we demonstrated that Cd and As were adsorbed on the sediments, in spite of the lowered pH due to CO2 dissolution in the groundwater. Cd concentrations in the effluent were below the Cd MCL, even for sediments without detectable carbonate to buffer the pH. Arsenic concentrations in the effluent were also significantly lower than the influent concentration, suggesting that the sediments tested have the capacity to mitigate the coupled adverse effects of CO2 leakage and brine intrusion. The mitigation capacity of sediment is a function of its geochemical properties (e.g., the presence of carbonate minerals, adsorbed As, and phosphate). other studies, including natural analogues,7,11,16,22 in situ CO2 injection in field sites,14,23−29 and laboratory column and batch studies12,13,30−36 have also been conducted to evaluate potential risks associated with CO2 leakage into groundwater aquifers. Wang and Jaffe were among the early researchers who used numerical simulation to predict trace metal (Pb) mobilization as a result of CO2 intrusion into hypothetical aquifers comprised of PbS/CaCO3 or PbS/quartz.19 They concluded that a CO2-induced decrease in pH would increase dissolution of PbS and the aqueous concentration of Pb. Based on their simulation results they also concluded that the presence of carbonate minerals can increase the buffering capacity of the aquifer, and thus mitigate the effect of CO2 intrusion on the mobilization of trace metals. In an analog study conducted in New Mexico, Keating et al. reported that despite relatively high levels of dissolved CO2, trace element mobility was not significant.16 They attributed this minor effect of CO2 intrusion to the high buffering capacity of the studied groundwater aquifer. In a column study, Frye et al. also observed that the geochemical properties of aquifer materials significantly influenced the response of groundwater quality to the intrusion of CO2.36 They found that calcite content as low as 10%

1. INTRODUCTION Although geologic CO2 sequestration (GCS) has been proposed as a viable option to reduce atmospheric CO2 emission,1−3 understandable concerns have been raised about possible impacts to shallow aquifers if CO2 were to leak from its storage reservoirs during or after GCS operations.4−6 Analog studies of geologic environments containing large amounts of CO2 have shown that leakage processes are possible in GCS.5 Lewicki et al. summarized CO2 leakage incidents from natural and industrial analogues revealing that in more than 20 occasions worldwide, CO2 escaped mainly through faults, fractures, and wells.7 Through these conduits, leaked CO2 as well as the brine from the storage reservoirs may intrude into groundwater aquifers, and the consequent biogeochemical reactions may cause the mobilization of trace metals or toxic organic compounds, thus causing the contamination of drinking water. Contamination of shallow groundwater with trace metals was the major concern for the current studies in evaluating the risks associated with CO2 leakage.8−14 Two sources must be considered when estimating the potential adverse impacts to an overlying aquifer from toxic metals resulting from CO2 leakage. The first source is the aquifer material itself. Intrusion of CO2 into shallow groundwater aquifer will decrease the pH of the groundwater, potentially dissolving aquifer minerals and releasing trace metals into the aqueous phase. Studies on this topic have recently increased. While numerical simulation was the major tool used to predict the effect of CO2 leakage,9,15−21 © XXXX American Chemical Society

Received: February 25, 2015 Revised: June 2, 2015 Accepted: June 3, 2015

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experiments. The SGW was prepared based on USGS data for the composition of the High Plains aquifer groundwater (SI Table S1 and S2).38 The final pH of the SGW was adjusted to ∼7.6 with 1 M KOH or 2% HNO3. Spiked-SGW was prepared by spiking the SGW with toxic trace metals, Cd and As, to simulate the scenario that both CO2 and brine containing toxic metals leak from a deep CO2 storage reservoir and intrude into the groundwater aquifer. Cd and As stock solutions were prepared by dissolving Na2HAsO4·7H2O and Cd(NO3)2 in water. Selection of As(V) for spiked-SGW instead of As(III) was because 1) previous research has shown As(V) is present in the deep saline aquifer under geological CO2 sequestration conditions,41 and 2) using As(V) instead of As(III) can avoid the complexity of potential As(III) oxidation during the experiment. The concentrations of Cd (40 μg/L) and As (114 μg/L) in the spiked-SGW were selected based on the predicted maximum concentrations of these metals in High Plains aquifer.42 While the intrusion of saline water into a groundwater aquifer may cause the ionic strength to increase, we did not adjust the ionic strength of the SGW or spiked SGW because (1) a previous study has reported that ionic strength increase has little influence on metal leaching;36 and (2) low ionic strength allows low dilution for analyses of trace metals and thus, lower detection limits. 2.3. Batch Experiments. A series of batch experiments were conducted to study major, minor, and trace element mobilization from sediments. The details for batch experiments have been described before (see SI).40 Briefly, 5 g sediment and 15 mL spiked-SGW were placed in the batch reactor (125 mL bottles). Before CO2 (99.998%) was introduced to the reactor, the sediment and the spiked-SGW were equilibrated in the reactor for ∼15 h. The flow rate of CO2 was 70 ± 10 mL/min. For each selected elapsed time, duplicate runs with continuous CO2 flow and one control experiment without CO2 were conducted simultaneously. At the end of each experiment, the reactor was opened and solution pH was measured immediately. An aliquot of aqueous sample was taken and prepared for elemental analyses with inductively coupled plasma-mass spectrometry (ICP-MS) and inductively coupled plasma-optical emission spectrometry (ICP-OES). The batch experiments were all conducted at room temperature (21 ± 2 °C). 2.4. Column Experiments. Polyvinyl chloride (PVC) columns with an inner diameter of 2.4 cm and length of 5.39 cm were packed with sediment samples as described before.43,44 Details of column experiments are summarized in SI Table S2. The columns were initially flushed with the SGW to achieve hydrological equilibrium (i.e., full saturation). To avoid the influence of oxygen, N2 gas was continuously purged in the bottle containing the influent solution to remove oxygen. The deaerated SGW was injected continuously into the columns from the bottom inlet using a KLOEHN V6 syringe pump at a flow rate of about 0.03 mL/min. After the columns reached full saturation, the deaerated SGW flowed through the column for about 3 days. After that the influent was switched from N2purged SGW to CO2 gas-saturated spiked-SGW. CO2 gas was continuously purged in the influent bottle at a rate of 0.5 mL/ min. The effluent pH and Eh were continuously measured online. A fraction collector (i.e., a syringe pump with eight ports) was used to collect the effluent into plastic vials. Each vial collected 3.6 mL (2 h) of effluent. The columns ran with CO2 saturated spiked-SGW were conducted for 15−20 days. During this time, two stop-flows that lasted 3 and 5 days,

mitigated the effect of pH reduction and resulted in zero Cd desorption from the Cd-laden material used in their study.36 The second potential source of metals that could contaminate the shallow groundwater aquifer is brine in the deep storage formation that may contains metals. Injection of CO2 will cause pH reduction of the brine, which may facilitate the release of metals originally present in formation rocks. Karamalidis et al. studied rocks−brine−CO2 interactions under GCS-relevant conditions with natural rock collected from eight different formations.10 They reported that 50−75% of their experiments showed Pb, As, Cr, and Cd concentrations in the brine above the respective maximum contamination levels (MCLs). If the brine were to leak from the reservoir and intrude into an overlying groundwater aquifer, it could contaminate the aquifer with the metals. Keating et al. discussed the possibility of brine leakage and provided evidence of upward comigration of CO2 and saline water along a fault in sedimentary rocks.16 It is noteworthy that holistic site characterization and careful site selection will avoid natural hazards, such as faults or fractures for GCS projects.2,3,37 However, to fully assess the risks associated with GCS, it is necessary to gain a better understanding of the effect of coleakage of CO2 and brine. To our knowledge, little work has been done to consider the second potential source of metals and to study the fate of leaked metals in groundwater aquifers. The objective of this research is to study the effects of the leakage of CO2 and CO2saturated, metal-contaminated fluids from CO2 storage reservoirs on the water quality in overlying groundwater aquifers. Specifically, we aim to study the fate of trace metals and the geochemical reactions that influence trace metal mobilization/immobilization in groundwater aquifers. Natural aquifer materials collected from the High Plains aquifer in Kansas were used in this work. Both batch and column experiments were conducted. Batch experiments were used to simulate sudden and relatively short-lived CO2 release, while column experiments were used to simulate gradual leaks of CO2.4 The results from this research will support site selection, risk assessment, policy-making, and public education efforts associated with geologic carbon sequestration. 2. Experimental Section. 2.1. Sediments. The sediment samples used in this research were from the High Plains Aquifer in Kansas (see the Supporting Information (SI) Figure S1 for maps). The High Plains aquifer is a heterogeneous, unconsolidated aquifer consisting of clay, silt, sand, and gravels.38,39 Three sediments were collected from well “CNG” at depths of 2.1−2.4, 17.7−18.6, and 33.2−33.8 m. They will be referred to as CNG-a, CNG-b, and CNG-c, respectively. One sediment was collected from well “CAL121” at a depth of 45.7−46.0 m and will be referred to as CAL hereafter. The sediment samples arrived as loose sediments and were sieved to collect the 98% in mass for all four sediments) for batch and column experiments. A series of characterization analyses were performed for these sediment samples. An 8 M acid (HNO3) extraction at 90 °C was conducted to determine the acid extractable elemental composition of the sediments as described previously.40,41 Soil texture was also determined (see SI for more details). The sand, silt, and clay fractions were dried in an oven at 60 °C and their mineral compositions were determined quantitatively with Xray diffraction (QXRD, see SI for more details). 2.2. Synthetic Groundwater. Synthetic groundwater (SGW) was used in this work for both batch and column B

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Figure 1. pH changes in the batch study for CNG-a and CNG-b (a), CNG-c and CAL (b). The solid red lines represent the pH of SGW (pH 7.62 in all batch experiments). Error bars for the experiments conducted with CO2 correspond to the standard deviation of the means of two measurements for duplicate experiments.

MCLs, and Fe, Mn, and Zn which are regulated with secondary MCLs.46 It is noteworthy that metal release due to CO2 dissolution in groundwater and the consequent pH perturbation would be much less than strong acid extraction. 3.2. Batch Experiment. 3.2.1. pH. Batch experiments without CO2 (the blank experiment) and with CO2 flowing through the reactor were performed to study the effects of a sudden and relatively short-lived CO2 release to groundwater aquifers. The pH of the aqueous phase in the blank experiment was generally consistent with the SGW (Figure 1). After CO2 was introduced into the batch reactor, a sharp drop in pH was observed for all four sediments, but the extent of pH drop and the evolution of pH as a function of time were different. CNGb was unique among the four sediments in the way it responded to the injected CO2. Only 4 h after the introduction of CO2, the pH in the CNG-b batch reactors dropped from 7.6 to 5.1 (Figure 1a). As the exposure time further increased, the pH slowly increased and eventually reached a steady-state at about 5.4 after 72 h, which is significantly lower than the steady-state pH (∼6.1) for the other three sediments. For these three sediments, the pH dropped from 7.6 to about 6.1 within 4 h and then reached a steady-state. The pH profile of the aqueous solution in the batch reactors is determined by both the rate of CO2 diffusion in the groundwater and the extent of rock dissolution.47 As CO2 diffuses and dissolves in the aqueous phase, the dissociation of carbonic acid causes a drop of solution pH. The pH drop will consequently cause the dissolution of rock minerals, which consumes hydrogen ions and causes the solution pH to increase. Because the rate of rock dissolution is generally much slower than the CO2 diffusion rate, the pH initially reaches a minimum, and then starts to increase as rock dissolution continues, following a trend similar to the one observed in Figure 1a for CNG-b. As for the other three sediments, it is likely that fast-dissolving minerals, such as carbonate minerals, were present and the dissolution of these minerals were nearly complete in 4 h. Two lines of evidence support this explanation. First, the results from sediment characterization indicated that calcite and/or dolomite were present in CNG-a, CNG-c, and CAL (SI Table S5 and S6). Second, significantly higher Ca and Mg concentrations for these three sediments than CNG-b were observed in the aqueous phase (SI, Section 2.1 and Figure S2). 3.2.2. Major Elements. The chemical composition of the SGW was significantly affected by CO2 introduction. After 14 days exposure to CO2, the concentrations of major elements, including Ca, Mg, Sr, and Si (SI Figure S2) were more than 45% higher than that in the control experiments. Among the

respectively, were applied to assess the effect of time-dependent reactions and processes on chemical element release into the aqueous phase when the fluid residence time was increased. At the end of the experiments selected effluent samples were analyzed by both ICP-OES and ICP-MS to determine the concentrations of major and trace elements. Additional samples were selected after influent switch and stop flow events. Selected samples were analyzed with ion chromatography (IC) for anions. To identify the speciation of As, some samples were passed through an As-speciation cartridge (MetalSoft Center, Pscataway, NJ)45 and prepared for ICP-MS analysis. This cartridge can effectively remove As(V), leaving only As(III) in the aqueous phase.

3. RESULTS AND DISCUSSION 3.1. Characterization of Sediments. Soil texture was determined for the four sediment samples used in this work (SI Table S4). Sediment CAL belongs to the “sand” soil texture class and contains the highest amount of sand (92.5%) and lowest amount of silt (4.2%) and clay (3.2%) among the four samples. CNG-a, which was collected from the upper part (7− 8′) of the well “CNG” contains the highest silt (8.9%) and clay (10.3%), which makes it a sandy loam. The other two sediments from deeper areas of Well CNG (CNG-b and CNG-c) are loamy sand. Quantitative XRD analyses (QXRD) were conducted for the clay and silt samples separated from the sediments (SI Table S5). The results from both QXRD and texture analysis indicated that while quartz was the major component for the four sandy sediments, varying amounts of feldspar, mica, kaolinite, and carbonate minerals were identified in these samples. Carbonate minerals (calcite and/or dolomite) were identified in all the samples except for CNG-b. Combining the results from soil texture analyses and carbonate concentrations in the silt and clay sections from QXRD, we estimated the amount of calcite and dolomite concentrations in the sediments (SI Table S6). The results show that CNG-a, CNG-c, and CAL contain 0.55, 3.8, and 1.9 wt % calcite, CNG-a and CAL contain 0.12 and 0.05 wt % dolomite, and CNG-b has no detectable carbonates. Acid extraction analyses were conducted to determine acid extractable elements of the aquifer sediments (SI Table S7). These elements may potentially be released from aquifer sediments when they are exposed to CO2 in the scenarios of CO2 leakage from storage reservoirs. The results shown in SI Table S7 indicated that the sediments used in this work contain varied amounts of metals that are of environmental concerns, such as Ba, Cr, Pb, and As which are regulated with primary C

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Figure 2. Concentrations of trace elements, Ba (a), Mn (b), Cd (c), and As (d) as a function of reaction time in batch study. Error bars for the experiments conducted with CO2 correspond to the standard deviation of the means of two measurements for duplicate experiments.

four sediments, release of Ca, Mg, and Sr from CNG-b was significantly lower than the other three sediments (see SI section 2.1 for more detailed discussions). It is noteworthy that CNG-b was also unique in the release of P. For CNG-a, CNGc, and CAL sediments, P was not detected in the aqueous phase; however, a significant amount of P was observed 4 h after CO2 introduction to the batch reactor containing CNG-b (SI Figure S2e). The concentration of P decreased slightly during 4−24 h, then remained relatively stable at about 1.1 mg/ L. As a chelating agent, phosphate has a strong affinity to mineral surfaces.48−51 Therefore, phosphate may affect the behavior of other metals/oxyanions in this system, such as As. In addition, acid extraction result showed that CNG-a, CNG-c, and CAL also contained detectable amount of P (SI Table S7). The fact that no detectable P was observed in the aqueous phase in contact with these sediments suggests that P was associated with less soluble minerals in these three sediments than in CNG-b. 3.2.3. Trace Elements. CO2 introduction into the groundwater-sediment system facilitated the release of some trace metals of environmental concern. Trace metals, including Ba, Mn, Cd, and As that are regulated by the U.S. EPA were detected in our batch experiments (Figure 2 and Figure S3 in the SI for 0−50 h result). The release of Ba was significantly enhanced by the introduction of CO2 for all four sediment (Figure 2a). In the CO2-experiements, Ba concentrations generally reached a steady-state after 1−3 days, but these concentrations remained well below the MCL (i.e., 2000 μg/ L).46 Researchers from several groups studied the release of trace metals from different sediments due to CO2 exposure, but similar to this work, none of them detected Ba concentration greater than the Ba MCL.30,33,52 Unlike most of the other metals that were released quickly in the first few days and then reached a steady-state, Mn shows a significant, time-dependent increasing trend after the sediments were exposed to the CO2

gas (Figure 2B). Mn concentrations measured during 4 h-14 d were all above its secondary MCL (i.e., 50 μg/L). Two trace metals, Cd and As, were purposely added into the SGW to test their behaviors in the groundwater aquifer in the scenario when both CO2 and the saline water containing toxic trace metals leak from CO2 storage reservoirs. The results from batch experiments indicate that their concentrations in both control and CO 2 experiments were well below their concentrations in the original spiked-SGW (40 μg/L Cd and 114 μg/L As) (Figure 2c and d, and Figure S3 c and d). Cd was not detected in the control experiments for all four sediments, even at t=0, suggesting that the sediments have the capacity to adsorb all Cd in the spiked-SGW within the 15 h equilibrium time period (the adsorbed Cd was 0.12 mg/kg soil for all four sediments). The introduction of CO2 into the batch reactor increased aqueous Cd concentrations, suggesting Cd desorption from sediments occurred. Similar results were found in the column study conducted by Frye et al. in the absence of carbonates.36 As the reaction proceeded, a portion of the released Cd was adsorbed again, which can be seen from the decrease in Cd concentrations during the time frame of 4 h - 2 d (Figure 2c and S2c). After day 2, Cd concentrations remained relatively stable for all four sediments. Nevertheless, Cd concentrations were all below its MCL (5 μg/L) after 24 h exposure to CO2 for all four sediments. Unlike Cd, which was completely removed by the sediments before CO2 was introduced, 14−45 μg/L of As were detected in the control experiments at t = 0 (Figure 2d and SI Figure 2d). These values account for 12−40% of the original As concentration (114 μg/L) in the spiked-SGW, suggesting the sediments have the capacity to adsorb As from the spikedSGW. In control experiments without CO2, As concentrations (shown as empty symbols with dashed lines) generally decreased with time. After 2−7 days, As was totally removed from the aqueous phase in the blank batch reactors containing CNG-a, CNG-b, and CAL sediments, respectively, whereas As D

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Figure 3. pH changes in the column study for CNG-a and CNG-b (a), CNG-c and CAL (b). The black dashed lines indicate the time for influent switch from SGW to CO2-saturated spiked-SWG. The blue dotted lines indicate the time for 1st and 2nd stop flow.

Figure 4. Concentrations of trace elements, Mn (a), Ba (b), and As (c and d) as a function of reaction time in column study for different sediments. The concentrations of Mn, Ba, and As in the spiked-SGW were 0.0, 3.5, and 114 μg/L, respectively.

concentrations for CNG-c were still near its MCL (10 μg/L) after 7 days. This result suggests that without the presence of CO2, more than 90% of As has been adsorbed by the sediments under our experimental conditions. The effects of CO2 introduction on the behavior of As in the four sediments were different. The adsorption of As onto CNG-a and CNG-c was facilitated by CO2 introduction at the early time, but As concentrations eventually reached the same level as corresponding blank experiments. The introduction of CO2 did not significantly affect the behavior of As for CAL, whereas it hindered the adsorption of As onto CNG-b. The behavior of As in the batch reactor containing CNG-b was not as expected. Previous reports revealed that the adsorption of As(V) onto mineral surfaces increased as pH decreased;53−55 however, for CNG-b, the pH decrease caused by CO 2 introduction (Figure 1a) resulted in decreased As adsorption. This result is also different from the findings by Little et al., who conducted a batch study with natural sediments and found that CO2 introduction decreased As concentrations in the groundwater.30 One plausible explanation for the unexpected behavior of As in the CNG-b batch experiment could be the presence of competitive adsorption between phosphate and arsenate [As(V)]. As discussed in section 3.2.2, phosphate was only detected in the aqueous phase for CNG-b. With a strong affinity for solid surfaces, phosphate can compete for surface sites with arsenate, thus hindering the adsorption of As. This phenomenon has been observed in previous research.48−50

3.3. Column Experiment. 3.3.1. pH. Column studies were conducted to simulate gradual CO2 leakage scenarios. Similar to batch experiments, the pH of the effluent from the four columns exhibited a sharp decrease after the influent was switched from SGW to CO2-saturated spiked-SGW (Figure 3). The pH profiles for CNG-a and CNG-c show a similar pattern: after the introduction of CO2-saturated spiked-SGW, the pH dropped from 8.0 to about 6.5. Although application of stopflows caused the pH to increase slightly, the effluent pH returned to ∼6.5 shortly after the flow resumed and then remained stable. The column packed with the CAL sediment showed a sharp pH drop to about 5.5 within 10 h, suggesting advective flow pathways had higher contribution to the effluent than diffusive flow in this early period. As diffusive flow pathways followed, the pH gradually increased to about 6.5 and then remained relatively steady. The pH profile for CNG-b was unique among the four column experiments. Except for the pH increase during the two stop-flow events (Figure 3a), pH continuously decreased to ∼5.3, then reached a relatively steady state. The increase of pH during the stop-flow events is likely due to mineral dissolution which consumes protons. The steady-state pH for the four columns was consistent with that in the batch experiments, i.e. pHCNG‑a ≈ pHCNG‑c ≈ pHCAL > pHCNG‑b (Figure 1). 3.3.2. Major Elements. As CO2-saturated spiked-SGW flowed through the columns, the pattern of release of major elements as a function of time differed between elements (Figure S4, see SI section 2.2 for more detailed discussion). E

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suggesting similar amounts of As(III) and As(V) were in the effluent. Importantly, As(III) was present in the effluent of the three columns containing sediments from CNG well. Because the spiked As was exclusively As(V), the potential sources of the As(III) could be (1) the As(III) that was in the original sediments and was released into the aqueous phase as a result of mineral dissolution and/or desorption, or (2) the reduction of As(V) in the inlet spiked-SGW. We ruled out the second possibility based on the Eh of the effluent (SI Figure S5). Throughout the column experiments, the effluent Eh was consistently above 320 mV. Although Eh increased for short periods of time, it resumed to its original range (320−360 mV). Therefore, the As(III) in the effluent came from the sediments. Although CNG-b contained the lowest As among the four sediments (SI Table S7), As(III) appeared first in CNG-b effluent, indicating that the affinity of As(III) in CNG-b to solid surfaces was weaker than that in the other sediments under our experimental conditions. Acid extraction results (SI Table S7) revealed that As was detected in the four sediments with concentrations ranging from 0.8 to 4.1 μg/g. Both As(III) and As(V) could be present in the sediments. When the CO2-saturated spiked-SGW flowed through the column, for a period of time, As(V) in the influent was all adsorbed onto the surfaces of minerals. Thus, no As was detected in the effluent in the early time periods of the column experiments, as can be seen for CNG-a, CNG-c, and CAL in Figure 4c−d. As more and more As(V) covered the surface sites, at some point, As(III) which has less affinity to solid surfaces than As(V) under acidic and near neutral conditions,53 can desorb due to the competitive adsorption of As(V). As a result, As(III) started to appear in the effluent and its concentration increased gradually, as we observed for CNG-a and CNG-c. The appearance of As(V), but no As(III) in the column CAL after the second stop-flow indicates that little As(III) was present in this sediment. The behavior of As in column CNG-b was unique, which is likely related to the phosphate detected in the effluent (SI Figure S4e). Previous studies have reported that phosphate competes for surface sites with both As(V) and As(III).48−50 Therefore, both As(III) and As(V) were detected in the effluent. However, it is noteworthy that within the time span of the column experiment, As(V) concentrations in the effluent were lower than that in the influent (114 μg/L), suggesting that part of the As in the influent was adsorbed on the sediments, despite the presence of phosphate. In our previous work, a batch experiment was conducted to study the release of trace metals from natural rock samples from the Eau Clair formation in the presence of CO2 under high pressure (10.1 MPa).41 Similar to this work, we found that both As(III) and As(V) were released into the aqueous solution, and we provided evidence that the release of As was due to the competitive adsorption between arsenate/arsenite and other anions, including phosphate. Although the experimental conditions are different between this work and our previous study, a similar mechanism, i.e. competitive adsorption, may occur in both reaction systems and affect the mobilization of As, which results in the earlier release of As from CNG-b than from the other three sediments (Figure 4c and d). At the end of the column experiments, total As concentrations were near or above the As MCL (10 μg/L).

Similar to the batch experiments, P was detected only in the effluent of the column packed with CNG-b (SI Figure S4e). IC analysis of effluent samples indicate that the phosphate concentration measured with IC was essentially the same as the total P measured with ICP-OES (SI Figure S4e), confirming that P was present in the effluent as phosphate. 3.3.3. Trace Elements. Some trace metals of environmental concern were detected in the effluent of the column study (Figure 4), but not all of the trace metals that were detected in the batch study were detected in the column experiments. Although Cd (40 μg/L) was present in the spiked-SGW for the column study, it was not detected in the effluent of any of the four columns throughout the experiments, suggesting that all the Cd in the influent was adsorbed on the sediments. Based on the volume of spiked-SGW flowed through the columns, we estimated the adsorbed Cd in the four sediments at the end of the column experiment was 0.68, 0.77, 0.78, and 0.64 mg/kg soil for CNG-a, CNG-b, CNG-c, and CAL, respectively. Among the metals regulated by U.S. EPA, only Ba, Mn, and As were detected in the effluents (Figure 4). After the inlet solution was switched to CO2-saturated spiked-SGW, Ba release was significantly enhanced. The Ba concentration peaked before the first stop-flow and then gradually decreased. For all four sediments, Ba concentrations in the effluents were well below its MCL (2000 μg/L), which is consistent with our observation in the batch study. Except for CNG-b, the release of Mn from the other sediments was not significantly influenced by the introduction of CO2-saturated SGW. Mn concentrations increased at later stages of the study and at the end of the column study the effluent Mn concentrations were all above its MCL (50 μg/L). The significantly higher Mn concentrations in CNG-b effluent is likely related to the low pH observed in this experiment, which resulted in the dissolution of Mn-containing minerals. Recently Kirsch et al. reported that carbonates could be the dominant source of mobile metals, including Mn, that are released to the groundwater in the scenarios of CO2 leakage;12 however, the different patterns of Mn, Ca, and Mg from this work suggest that the release of Mn from the High Plains aquifer sediments was not from carbonates. Arsenic was one of the two trace metals (As and Cd) added in the spiked-SGW. Our batch study revealed that the High Plains sediments have the capacity to adsorb As. Therefore, it was expected that As would be released in the columns effluents later than the other metals, such as Ba and Mn. The profiles of As release shown in Figure 4c and d indicate that, as expected, As was not detected in the effluent of columns packed with CNG-a, CNG-c, and CAL sediments until after the second stop-flow; however, the breakthrough of As for CNG-b occurred within 20 h after CO2-saturated spiked-SGW was introduced to the column, which was much earlier than the other three columns, suggesting processes other than adsorption influenced the behavior of As. Speciation analyses for As in the effluent suggest that part of the As in the effluent was from the sediment, not the spiked As. The added As was exclusively arsenate [As(V),114 μg/L]; however, arsenite [As(III)] was detected in the effluent from some columns. Figure 4c and d shows that the As(III) concentrations in the effluent of CNG-a and CNG-c columns were essentially the same as the total As, suggesting arsenate [As(V)] was negligible in the effluent. For CAL, As(III) was not detected, indicating that the released As was exclusively As(V). For CNG-b, As(III) accounts for 45−56% of total As, F

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Environmental Science & Technology

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4. ENVIRONMENTAL IMPLICATION Although extensive research has been conducted to study metal release in the scenario of CO2 leakage, to the best of our knowledge, no experimental study has been conducted to evaluate the behaviors of trace metals leaked from the storage reservoir together with CO2. With natural sediments collected from High Plains sandy aquifer, we demonstrated that the natural sediments of this groundwater aquifers have the capacity to mitigate the adverse effect of the leakage of CO2 and the concomitant leakage of brine containing toxic trace metals, such as Cd and As. The amount of carbonates available for buffering of pH in shallow groundwater has been suggested as an important factor in GCS site selection.11,30,36 Frye et al. suggested that calcite content as low as 10% can mitigate the adverse effects of CO2 leakage into groundwater aquifers.36 However, the results from this work showed that Cd and As that entered the shallow groundwater aquifer from deep storage formation can be adsorbed by the aquifer sediments containing only 0−4% carbonates. In addition, sediments containing as low as 0.67% carbonates (CNG-a) can maintain the pH at 6.5 which is EPA’s minimum MCL for pH.46 Therefore, deep saline formations with overlying sandy groundwater aquifers which usually contain low carbonates, such as High Plains aquifer, should not be excluded as potential GCS sites. The mitigation capacity of the groundwater aquifer needs to be considered in the integrated risk assessment modeling.

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ASSOCIATED CONTENT

S Supporting Information *

Descriptive texts, figures, and tables of experimental procedures and data analysis.The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/ acs.est.5b01004.

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AUTHOR INFORMATION

Corresponding Author

*Phone: (217) 300-3157; fax: (217) 244-7004; e-mail: [email protected]. Present Address †

Illinois State Geological Survey, 615 East Peabody Drive, MC650, Champaign, Illinois 61820, United States. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS Funding for this research was provided by the National Risk Assessment Partnership (NRAP) in the U.S. DOE Office of Fossil Energy under DOE Contract Number DE AC05 76RL01830. XRD analyses were performed in the Environmental Molecular Sciences Laboratory (EMSL), a national scientific user facility sponsored by the Department of Energy’s Office of Biological and Environmental Research and located at PNNL.

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REFERENCES

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DOI: 10.1021/acs.est.5b01004 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

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DOI: 10.1021/acs.est.5b01004 Environ. Sci. Technol. XXXX, XXX, XXX−XXX