Mercury Mobilization in a Flooded Soil by Incorporation into Metallic

Article pubs.acs.org/est

Mercury Mobilization in a Flooded Soil by Incorporation into Metallic Copper and Metal Sulfide Nanoparticles Anke F. Hofacker,‡ Andreas Voegelin,§,* Ralf Kaegi,§ and Ruben Kretzschmar‡ ‡

Soil Chemistry Group, Institute of Biogeochemistry and Pollutant Dynamics, ETH Zurich, CHN, CH-8092 Zurich, Switzerland Eawag, Swiss Federal Institute of Aquatic Science and Technology, CH-8600 Dübendorf, Switzerland

§

S Supporting Information *

ABSTRACT: Mercury is a highly toxic priority pollutant that can be released from wetlands as a result of biogeochemical redox processes. To investigate the temperature-dependent release of colloidal and dissolved Hg induced by flooding of a contaminated riparian soil, we performed laboratory microcosm experiments at 5, 14, and 23 °C. Our results demonstrate substantial colloidal Hg mobilization concomitant with Cu prior to the main period of sulfate reduction. For Cu, we previously showed that this mobilization was due to biomineralization of metallic Cu nanoparticles associated with suspended bacteria. X-ray absorption spectroscopy at the Hg LIII-edge showed that colloidal Hg corresponded to Hg substituting for Cu in the metallic Cu nanoparticles. Over the course of microbial sulfate reduction, colloidal Hg concentrations decreased but continued to dominate total Hg in the pore water for up to 5 weeks of flooding at all temperatures. Transmission electron microscopy (TEM) suggested that Hg became associated with Cu-rich mixed metal sulfide nanoparticles. The formation of Hg-containing metallic Cu and metal sulfide nanoparticles in contaminated riparian soils may influence the availability of Hg for methylation or volatilization processes and has substantial potential to drive Hg release into adjacent water bodies.

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Hg(II).13,14 Microbial Hg(II) reduction to Hg(0) is mostly due to bacterial detoxification catalyzed by a mercury reductase (MerA).15 This pathway was demonstrated to occur mainly in aerobic and only to a lesser extent in anaerobic environments.16,17 In addition, dissimilatory metal reducing bacteria (DMRB) were reported to reduce Hg(II) via a merindependent pathway.18 As a soft metal cation, Hg2+ not only exhibits a very high binding affinity for thiol-groups and disulfide in natural organic matter (NOM),19 but also forms extremely stable Hg-sulfide minerals.20 Therefore, under suboxic to anoxic conditions, organic thiol-groups and sulfide formed by sulfate-reducing bacteria (SRB) compete for Hg complexation21,22 and may both reduce the extent of CH3Hg+ or Hg(0) formation.12,23 In sulfidic soils and sediments, metastable metacinnabar (β-HgS) rather than thermodynamically more stable cinnabar (α-HgS) is expected to form24 via amorphous HgS.25 Submicrometer-sized metacinnabar crystals have been identified in a highly Hg contaminated drained riparian soil and were interpreted as a remnant of a past flooding event.26

INTRODUCTION Mercury (Hg) is a global high-priority pollutant. Microbially formed methylmercury (CH3Hg+)1−5 is extremely neurotoxic and poses a serious threat due to its pronounced tendency to bioaccumulate along the food chain. Elemental Hg(0) is volatile and contributes to long-range atmospheric Hg transport, making Hg a global pollutant that also affects remote areas.6 Riparian and coastal wetlands are important sinks for anthropogenic contaminants including Hg. Consequently, they may also release accumulated contaminants into surface and groundwater bodies or to the atmosphere. Wetland soils and sediments experience periodic flooding and associated changes in soil redox conditions that strongly affect the speciation and mobility of metal contaminants like Hg, which occurs in the redox states Hg(II), Hg(I), and Hg(0). Suboxic to anoxic conditions may induce the formation of Hg(0) in wetland soils.7 Reduction of Hg(II) to Hg(0) proceeds readily via abiotic and biotic pathways, making wetland soils an important source of Hg emissions to the atmosphere.6,8,9 Humic acids (HA) can reduce Hg(II) to Hg(0) under oxic and anoxic conditions,10−12 but may also stabilize Hg(II) at low concentrations through strong complexation by functional groups containing reduced sulfur.12 In analogy, binding of Hg(II) to high-affinity sulfhydryl-groups on Bacillus subtilis was observed to diminish Hg(II) reduction by green rust and magnetite, which are known to effectively reduce © 2013 American Chemical Society

Received: Revised: Accepted: Published: 7739

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contaminated with 1.26 and 0.91 mg/kg Hg and 279 and 160 mg/kg Cu, respectively; concentrations clearly above average contents of 0.11 mg/kg Hg (n = 696) and 17 mg/kg Cu (n = 783) reported for floodplain soils in Europe.41 Soil pH, texture, and total element contents are reported in Table S1 in the Supporting Information (SI). Soil Microcosm Experiments. Soil flooding experiments were conducted in a series of microcosms (one per data point) with air-dried soil that was flooded with synthetic river water and incubated for periods of 0.1 to 33 days in a temperaturecontrolled incubator (Binder KB 400, Germany; ±0.1 °C) in the dark. Incubations were conducted at 5 and 14 °C with unamended soil I (experiment A) and at 23 °C with lactateamended soil II (experiment B). Experiment A (unamended soil I; 5 and 14 °C) was performed with the same soil and microcosm setup as used in our previous studies (Figure S1).20,36,37 450 g of soil I (200− 2000 μm) were equilibrated with 1500 mL of synthetic river water containing 0.6 mM NaCl, 0.6 mM CaSO4, and 0.3 mM Mg(NO3)2 on an end-over-end shaker for two hours at room temperature. After centrifugation at 600 g for 15 min, the wet soil was placed into the microcosm and flooded with 500 mL of synthetic river water. Microcosm geometry consisted of a water-saturated soil layer of ∼9 cm height (pore water-to-soil ratio: 1.0 L/kg) submerged under a 6 cm layer of supernatant floodwater. Each microcosm was equipped with an open-pore suction cup (10−16 μm pore size; EcoTech, Germany) mounted ∼3.5 cm below the soil-supernatant interface and was open to air via a 3-mm hole in the lid. Experiment B (lactate-amended soil II; 23 °C) was performed analogously to experiment A. Soil microbial activity was stimulated by the higher incubation temperature and a lactate amendment. Lactate has been reported to promote the activity of different groups of soil microorganisms including Fe and sulfate reducers42 and was therefore considered suitable to stimulate soil reduction. Preliminary tests confirmed higher colloidal Hg formation in the presence of lactate, which allowed accumulation of sufficient colloidal material for Hg speciation by XAS. 550 g of dry soil (soil II, 200−2000 μm) was placed into each microcosm, flooded with 700 mL of synthetic river water amended with 12 mM lactate, and gently mixed to remove enclosed air. The water-saturated soil layer was ∼10 cm high (pore water-to-soil ratio: 1.27 L/kg) and submerged under a 4-cm layer of supernatant floodwater. The suction cup was mounted 4.5 cm below the soil-supernatant interface. Pore Water and Colloid Sampling. Soil pore water was directly withdrawn into an anoxic glovebox (MBraun MB 200B, M. Braun, Germany, pO2 Cd ∼ Pb > Zn > Fe) and the kinetics of metal release and diffusion to the reaction zone.20 Consequently, different metals may coprecipitate in the pore water depending on the local metal and sulfide concentrations. Furthermore, coprecipitated Fe or Zn may be exchanged for Hg, Ag, Cu, Cd, or Pb,48 due to the lower stability of Zn and Fe monosulfides. Hence, with respect to Hg, we would expect HgS to form first at very low dissolved sulfide concentrations, unless Hg supply was limited by kinetic factors. In the case of Cu, we previously attributed the decrease in colloidal Cu to the onset of metal sulfide precipitation.37 Thus, at 14 °C, the earlier decrease of colloidal Hg than Cu may indeed reflect initial formation of Hgrich sulfide colloids. Note however that even very small fractions of Cu or other trace metals coprecipitating with Hg are largely sufficient to render Hg a minor element in the resulting mixed metal sulfide precipitates. At 23 °C, early Hgsulfide precipitation may have been reflected in the more pronounced decrease of colloidal Hg relative to Cu after day 3. At 5 °C, on the other hand, slow sulfide generation may have allowed for the initial formation of colloids with higher Hg fractions because there was more time available for Hg supply and reaction with biogenic sulfide. Whether the Hg within Hg-

rich metal sulfide nanoparticles was homogeneously distributed at the atomic level or localized in (sub)nanometric HgS clusters could not be resolved by our TEM analyses. Incorporation of Hg(0) into Metallic Cu Nanoparticles. The Hg LIII-edge XANES and EXAFS spectra of colloids collected after 3 days of soil flooding at 23 °C (lactate-amended soil) and of selected reference spectra (including Cu metal measured at the Cu K-edge) are shown in Figure 3. The sample XANES and EXAFS spectra did not exhibit any of the spectral features characteristic of α-HgS, β-HgS, or elemental Hg(0), suggesting that none of these phases dominated colloidal Hg speciation after 3 days of soil flooding. Interestingly, however, the oscillations of the EXAFS spectrum exhibited remarkable similarity to the Cu K-edge EXAFS spectrum of metallic Cu (Figure 3). This similarity was also apparent in the Fouriertransformed EXAFS spectra of colloidal Hg and metallic Cu (Figure 3d), where peaks at similar distances were observed out to 8 Å, albeit with a shifted real part due to the difference in the absorber phase shift. This striking similarity strongly suggested that Hg substituted for Cu in the structure of metallic Cu. Considering the low molar Hg/Cu ratio of the studied sample of ∼0.002 (Figure S3), we assumed Hg to be exclusively surrounded by Cu atoms (no Hg−Hg neighbors) and, accordingly, fit the spectrum based on theoretical scattering paths calculated from the structure of metallic Cu with Hg substituting the absorbing atom (Figure 3d). Based on a fit model including the first 3 single scattering paths and focused multiple scattering paths at distances of ∼5.1 and 7.6 Å with degeneracies fixed to their theoretical values, the Fouriertransformed spectrum could be reproduced up to R+ΔR = 8 Å (Figure 3d) with fitted half-path lengths in line with Hg substituting in the Cu metal structure (Table 1). The first-shell Hg−Cu distance of 2.62 Å was slightly larger than the respective Cu−Cu distance in Cu metal (2.55 Å), whereas the second- and third-shell Hg−Cu distances did not significantly deviate from the respective Cu−Cu distances, reflecting the similar covalent radii of Hg and Cu (both 1.32 Å according to Cordero et al.49). The fact that multiple scattering out to 7.6 Å contributed to the Hg EXAFS signal was in line with earlier Cu 7743

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Figure 3. Hg LIII-edge XANES (a) and EXAFS (b) spectra and Fourier-transform magnitude (c) of colloidal Hg collected after 3 days of soil flooding at 23 °C compared to reference spectra of α-HgS (cinnabar), β-HgS (metacinnabar), elemental Hg(0) and metallic Cu(0) (Cu K-edge spectrum; EXAFS not corrected for phase shift). The Hg and Cu XANES spectra are plotted relative to E0 of 12 286 and 8981 eV, respectively. The 3-day colloid spectrum is the average of two nearly identical spectra of different 3-day samples (shown in Figure S4). In panel (d), the Fouriertransform magnitude and real part of the Hg colloid EXAFS spectrum are compared to the metallic Cu spectrum and a shell-fit for Hg substituted in metallic Cu (dotted line; fit parameters in Table 1).

Au−Ag alloy nanoparticles by bacteria and fungi,45,46 of Au nanoparticles and nuggets by bacteria,53−56 and of Au−Ag−Cu nanoparticles by plants57 has previously been observed. Our study provides first evidence for the formation of a Cu−Hg alloy nanoparticles driven by microbial Cu metal biomineralization.37 On the other hand, Cu amalgamation with Hg is wellknown from ore deposits, where the minerals kolymite (Cu7Hg6) and its dimorph belendorffite were described 20− 30 years ago.58−60 In the region of Lake Superior, strong correlations of Cu with Hg were observed within ore deposits and shoreline lake sediments,61,62 and several Hg-rich Cu and Ag ores occur. Interestingly, in some EDX spectra of metallic Cu nanoparticles, we also identified Ag (Figure S5), indicating that metals other than Hg may be colloidally comobilized by bacterial Cu metal mineralization.

XAS and TEM results showing that the Cu metal nanoparticles were well-crystallized.37 Due to the low Hg content in metallic Cu nanoparticles, we were not able to detect Hg by TEM-EDX. However, TEM-EDX revealed the association of Ag with metallic Cu nanoparticles (Figure S5), probably due to structural substitution in analogy to Hg. Table 1. Structural Parameters Determined by Shell Fitting of the EXAFS Spectrum of Colloidal Hg after 3 Days of Floodinga path

degb

Hg−Cu Hg−Cu Hg−Cu ms1/ms2/ms3e ms4/ms5f

12 6 24 24/12/12 24/24

R [Å]c

σ2 [Å2]d

2.62 3.58 4.43 5.16 7.52

0.006 0.009 0.009 0.009 0.005

(1) (3) (2) (1) (4)

(1) (3) (2) (1) (3)

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ENVIRONMENTAL IMPLICATIONS This study was conducted with two topsoil samples from the contaminated floodplain of the river Mulde in Germany, which contained 0.91−1.26 mg/kg Hg and 160−279 mg/kg Cu. These contents correspond to about ten times the average contents of European floodplain soils (0.11 mg/kg Hg; 17 mg/ kg Cu)41 and fall into the range reported for contaminated floodplains along European rivers, including the river Elbe and its tributaries Mulde and Saale.8,38,63 In recent work using the same soil as well as a Cu-spiked paddy soil from Bangladesh, we showed that the formation of metallic Cu nanoparticles may be a common process induced by soil flooding before sulfate reducing conditions are reached or when sulfide formation is limited by low sulfate availability.20,36,64 Our new results show that under the same conditions, formation of Hg-bearing metallic Cu nanoparticles may represent an important Hg transformation pathway in soils. Considering the existence of Cu7Hg6 minerals,58−60 formation of Hg−Cu amalgam nanoparticles may also be relevant at much higher soil Hg/Cu ratios than studied here. During sulfate reduction, Hg is effectively incorporated into nanoparticulate metal sulfides that may either

The shell fit of the k3-weighted spectrum was performed in R-space over the R + ΔR-range 1.5−8.0 Å. The passive amplitude reduction factor was 0.91 (±0.10). The energy-shift was 1.6 (±1.2) eV, and the goodness of fit indicated by the normalized sum of squared residuals (NSSR = ∑(datai − fiti)2/∑(datai)2) equaled 0.0004. Parameter uncertainties are given in parentheses for the last significant digit. b Path degeneracy (coordination number for single-scattering paths). c Average half path length (interatomic distance for single scattering paths). dDebye−Waller factor. eFocused collinear multiple scattering paths, ms1: Hg−Cu−Cu, ms2: Hg−Cu−Hg-Cu, ms3: Hg−Cu−CuCu. fms4: Hg−Cu−Hg-Cu, ms5: Hg−Cu−Cu-Cu. a

The reduction of Hg(II) to Hg(0) and incorporation into the metallic Cu may involve initial Hg(II) reduction by mixed Fe2+/Fe3+ (hydr)oxides,13,14 reduced humic acids,10−12 or dissimilatory metal reducing bacteria.18 Direct Hg reduction at the metallic Cu surface,50 however, represents the most plausible pathway of Hg reduction and incorporation during Cu crystal growth. Effective Hg reduction and uptake by Ag nanoparticles has recently been reported,51,52 with smaller nanoparticles incorporating relatively more Hg.52 Formation of 7744

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from contaminated floodplain soils to the atmosphere with simple field measurement techniques. Water Air Soil Pollut. 2002, 135, 39−54. (7) Hsu-Kim, H.; Kucharzyk, K. H.; Zhang, T.; Deshusses, M. A. Mechanisms regulating mercury bioavailability for methylating microorganisms in the aquatic environment: a critical review. Environ. Sci. Technol. 2013, 47, 2441−2456. (8) During, A.; Rinklebe, J.; Bohme, F.; Wennrich, R.; Stark, H. J.; Mothes, S.; Du Laing, G.; Schulz, E.; Neue, H. U. Mercury volatilization from three floodplain soils at the central Elbe River, Germany. Soil. Sediment. Contam. 2009, 18, 429−444. (9) Fitzgerald, W. F.; Lamborg, C. H.; Hammerschmidt, C. R. Marine biogeochemical cycling of mercury. Chem. Rev. 2007, 107, 641−662. (10) Alberts, J. J.; Schindler, J. E.; Miller, R. W.; Nutter, D. E. Elemental mercury evolution mediated by humic acid. Science 1974, 184, 895−896. (11) Allard, B.; Arsenie, I. Abiotic reduction of mercury by humic substances in aquatic system - an important process for the mercury cycle. Water Air Soil Pollut. 1991, 56, 457−464. (12) Gu, B. H.; Bian, Y. R.; Miller, C. L.; Dong, W. M.; Jiang, X.; Liang, L. Y. Mercury reduction and complexation by natural organic matter in anoxic environments. Proc. Natl. Acad. Sci. U.S.A. 2011, 108, 1479−1483. (13) O’Loughlin, E. J.; Kelly, S. D.; Kemner, K. M.; Csencsits, R.; Cook, R. E. Reduction of Ag(I), Au(III), Cu(II), and Hg(II) by Fe(II)/Fe(III) hydroxysulfate green rust. Chemosphere 2003, 53, 437− 446. (14) Wiatrowski, H. A.; Das, S.; Kukkadapu, R.; Ilton, E. S.; Barkay, T.; Yee, N. Reduction of Hg(II) to Hg(0) by magnetite. Environ. Sci. Technol. 2009, 43, 5307−5313. (15) Barkay, T.; Miller, S. M.; Summers, A. O. Bacterial mercury resistance from atoms to ecosystems. FEMS Microbiol. Rev. 2003, 27, 355−384. (16) Schaefer, J. K.; Letowski, J.; Barkay, T. mer-Mediated resistance and volatilization of Hg(II) under anaerobic conditions. Geomicrobiol. J. 2002, 19, 87−102. (17) Barkay, T.; Kritee, K.; Boyd, E.; Geesey, G. A thermophilic bacterial origin and subsequent constraints by redox, light and salinity on the evolution of the microbial mercuric reductase. Environ. Microbiol. 2010, 12, 2904−2917. (18) Wiatrowski, H. A.; Ward, P. M.; Barkay, T. Novel reduction of mercury(II) by mercury-sensitive dissimilatory metal reducing bacteria. Environ. Sci. Technol. 2006, 40, 6690−6696. (19) Xia, K.; Skyllberg, U. L.; Bleam, W. F.; Bloom, P. R.; Nater, E. A.; Helmke, P. A. X-ray absorption spectroscopic evidence for the complexation of Hg(II) by reduced sulfur in soil humic substances. Environ. Sci. Technol. 1999, 33, 257−261. (20) Weber, F.-A.; Voegelin, A.; Kretzschmar, R. Multi-metal contaminant dynamics in temporarily flooded soil under sulfate limitation. Geochim. Cosmochim. Acta 2009, 73, 5513−5527. (21) Skyllberg, U. Competition among thiols and inorganic sulfides and polysulfides for Hg and MeHg in wetland soils and sediments under suboxic conditions: illumination of controversies and implications for MeHg net production. J. Geophys. Res. - Biogeosciences 2008, 113. (22) Skyllberg, U.; Drott, A. Competition between disordered iron sulfide and natural organic matter associated thiols for mercury(II) - an EXAFS study. Environ. Sci. Technol. 2010, 44, 1254−1259. (23) Benoit, J. M.; Gilmour, C. C.; Mason, R. P.; Heyes, A. Sulfide controls on mercury speciation and bioavailability to methylating bacteria in sediment pore waters. Environ. Sci. Technol. 1999, 33, 951− 957. (24) Slowey, A. J. Rate of formation and dissolution of mercury sulfide nanoparticles: the dual role of natural organic matter. Geochim. Cosmochim. Acta 2010, 74, 4693−4708. (25) Charnock, J. M.; Moyes, L. N.; Pattrick, R. A. D.; Mosselmans, J. F. W.; Vaughan, D. J.; Livens, F. R. The structural evolution of mercury sulfide precipitate: an XAS and XRD study. Am. Mineral. 2003, 88, 1197−1203.

form by transformation of Cu(0) nanoparticles or by precipitation from solution. Based on the extremely high thermodynamic stability of HgS, higher soil Hg/Cu ratios are expected to favor pure HgS over mixed metal sulfide formation. If quantitatively relevant in the soil matrix, both transformation mechanisms identified in the present work may affect the extent of Hg methylation and volatilization. In any case, this study shows that Hg incorporation into metallic Cu and metal sulfide nanoparticles can cause substantial colloidal Hg mobilization into the pore water. When exposed to molecular oxygen, Hgcontaining metallic Cu nanoparticles are expected to dissolve within hours to days,64,65 whereas metal sulfide nanoparticles can be stable against oxidation in oxygenated water up to several weeks.26,33,64 Thus, Hg-containing metal sulfide nanoparticles may significantly contribute to long-range Hg transport upon their release into oxic aquifers or surface waters.

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ASSOCIATED CONTENT

S Supporting Information *

Soil properties and element contents (Table S1). Experimental microcosm setup (Figure S1). Temporal trends of pH, redox potential, DOC, Mn, Fe, sulfate (Figure S2). Colloidal, dissolved and soil Hg/Cu mole ratios (Figure S3). XAS of colloidal Hg of two separate samples (Figure S4). TEM-EDX results for metallic Cu nanoparticles (Figure S5). This material is available free of charge via the Internet at http://pubs.acs.org.

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AUTHOR INFORMATION

Corresponding Author

*E-mail [email protected], phone +41 58 765 5470, fax +41 58 765 5210. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS We acknowledge Kurt Barmettler for his support in the laboratory and Beate Fulda for her extensive help with XAS measurements. We thank Sergey Nikitenko and Miguel Silveira for assistance at the DUBBLE beamline (ESRF, Grenoble, France) and acknowledge DUBBLE and the Angströmquelle Karlsruhe (ANKA, Germany) for the allocation of beamtime. The electron microscopy centre at ETH Zurich (EMEZ) is acknowledged for providing access to their TEM instruments. The study was financially supported by ETH Zurich and Eawag.

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REFERENCES

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