1 Lecture 11. Global biogeochemical cycles of nitrogen and oxygen. Objectives: 1. Global biogeochemical cycle of nitrogen. 2. Global biogeochemical cycle of oxygen.
Lecture 12 – Global Biogeochemical Cycles (1) If rivers are the chief source of the dissolved salts in seawater, why is seawater not simply a concentrated version of average composition of all rivers?
The University of Maine [email protected] Marine Sciences Faculty Scholarship School of Marine Sciences 8-1-2009 Observing Biogeochemical Cycles at Global Scales
synthesis and characterization of vanadium oxide nanomaterials by tyler martin fears a dissertation presented to the graduate faculty of the missouri university of ...
A comparison of the effect of temperature on the reduction of N2 by purified molybdenum nitrogenase and vanadium nitrogenase of Azotobacter chroococcum showed differences in behaviour.
Unlike other RFBs, vanadium redox flow batteries (VRBs) use only one element (vanadium) in both tanks, exploiting vanadium’s ability to exist in several states. By using one element in both tanks, VRBs can overcome cross-contamination degradation, a
The global marine phosphorus cycle: ... increased concentrations of soluble P may offset the drop ... Reactive phosphorus in the water column is present under
SALINITY AND THE GLOBAL WATER CYCLE R. W. Schmitt (1), T. Boyer (2) ... surface salinity has much to tell us about the global ... The units are centimeters per year. Figure 1.b (bottom). Average surface salinity of the world ocean, contoured from the
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Global biogeochemical cycle of vanadium William H. Schlesingera,1, Emily M. Kleina, and Avner Vengosha a
Earth and Ocean Sciences, Nicholas School of the Environment, Duke University, Durham, NC 27708
Contributed by William H. Schlesinger, November 9, 2017 (sent for review September 1, 2017; reviewed by Robert A. Duce, Andrew J. Friedland, and James N. Galloway)
Synthesizing published data, we provide a quantitative summary of the global biogeochemical cycle of vanadium (V), including both human-derived and natural fluxes. Through mining of V ores (130 × 109 g V/y) and extraction and combustion of fossil fuels (600 × 109 g V/y), humans are the predominant force in the geochemical cycle of V at Earth’s surface. Human emissions of V to the atmosphere are now likely to exceed background emissions by as much as a factor of 1.7, and, presumably, we have altered the deposition of V from the atmosphere by a similar amount. Excessive V in air and water has potential, but poorly documented, consequences for human health. Much of the atmospheric flux probably derives from emissions from the combustion of fossil fuels, but the magnitude of this flux depends on the type of fuel, with relatively low emissions from coal and higher contributions from heavy crude oils, tar sands bitumen, and petroleum coke. Increasing interest in petroleum derived from unconventional deposits is likely to lead to greater emissions of V to the atmosphere in the near future. Our analysis further suggests that the flux of V in rivers has been incremented by about 15% from human activities. Overall, the budget of dissolved V in the oceans is remarkably well balanced—with about 40 × 109 g V/y to 50 × 109 g V/y inputs and outputs, and a mean residence time for dissolved V in seawater of about 130,000 y with respect to inputs from rivers. vanadium
| petroleum | geochemical cycle | aerosols | rock weathering
anadium (V) occurs in a wide range of earth materials and is a relatively abundant trace metal, with an average concentration in the upper continental crust (97 mg/kg) more than double those of nickel (Ni) and copper (Cu) (1). In modern society, the majority of V is used to improve the strength and corrosion resistance of steel; it is also of increasing strategic and technological interest as a specialty metal in electronics and batteries. V is an essential trace element in prokaryotic biochemistry, where it is found as an alternative to molybdenum in the molecular structure of nitrogenase, the enzyme of N fixation (2–4). It also appears in the structure of enzymes in the marine algae responsible for the formation of bromoform (5) and methyl bromide (6), which contributes to the depletion of stratospheric ozone. Whether V is an essential trace element in higher plants and animals is unknown (7), with some evidence favoring an essential role in at least some higher organisms (e.g., refs. 8 and 9). Higher plants accumulate about 1 mg/kg in their tissues (10). The mean concentration of V in the continental crust is about 97 mg/kg (1, 11), and ∼20 × 109 g V/y enters biogeochemical cycles at Earth’s surface through chemical weathering (12). V, which can exist in three common oxidation states, is largely found as the vanadate ion (H2VO4−) in natural oxidized waters of near-neutral pH. The dissolved concentration in river water is about 0.7 μg/L, less than half of the concentration of ∼1.8 μg/L in seawater (13). V is widely recognized as a trace contaminant in coal and oil and is released to the atmosphere through particle emissions associated with their combustion (14–17). While V has not received as much attention as other anthropogenically mobilized elements, data presented here show that the human perturbation of V cycling through the atmosphere may exceed that of mercury (Hg) and a variety of other metals (Table 1). V is not viewed as a E11092–E11100 | PNAS | Published online December 11, 2017
problematic environmental contaminant, but high concentrations of V can be toxic to humans and other organisms (18, 19). Reflecting a new level of concern, the State of California has recently imposed a new standard (15 μg/L) for V in drinking water (https://oehha.ca.gov/water/notification-level/proposed-notification-level-vanadium). In some regions, the release of V to the atmosphere and its deposition in natural ecosystems have declined in recent decades due to changes in fuel use and industrial practices (20, 21). Vanadium’s primary commercial use is in the manufacture of steel alloys to increase hardness and corrosion resistance. An emerging application is the development of V redox batteries as an alternative for lithium-ion batteries for storage of renewable energy. Worldwide, about 82 × 109 g of V are produced from virgin ore annually, and, at current rates of consumption, reserves (∼15 × 1012 g) are expected to meet demand into the next century (22). In the United States, recycling of products and fly ash provides as much as 40% of the V in commercial use, including exports (23). Huang et al. (13) provide a comprehensive review of the oxidation and reduction processes that transform V in Earth’s surface environments, and Hope (24) quantifies the movement of V on the present-day and future Earth, based on its mobilization by human activities. In this paper, we provide an updated, quantitative summary of the global biogeochemical cycle of V, including both human and natural fluxes. Using published data, we compile and calculate global-scale estimates of the flux of V in nature, especially as modified by recent changes in the combustion of coal and petroleum products and byproducts. Due to limited data, many of these estimates must be presented without boundaries of uncertainty, but we are encouraged by the general internal consistency that emerges when the independent flux estimates produce a roughly balanced global budget. Mobilization of V from Earth’s Crust Natural Sources. Natural V mobilization from Earth’s crust occurs dominantly through mechanical and chemical rock weathering, with lesser contributions from other pathways such as volcanic emissions and aeolian processes. For mechanical weathering, Hope (24) estimates that 2,300 × 109 g V/y is transported to the sea in suspended sediments. We suggest a somewhat (30%) Significance Human emissions of vanadium to atmosphere exceed natural sources by a factor of 1.7 and are destined to rise dramatically as we switch to the use of heavy oils, tar sands, and bitumen as combustion sources. Breathing vanadium-rich aerosols has unknown but potentially adverse health impacts. The human impacts on the global vanadium cycle parallel impacts on the global cycles for Pb and Hg. Author contributions: W.H.S., E.M.K., and A.V. designed research, performed research, analyzed data, and wrote the paper. Reviewers: R.A.D., Texas A&M University, retired; A.J.F., Dartmouth College; and J.N.G., University of Virginia. The authors declare no conflict of interest. Published under the PNAS license. 1
Table 1. Movement of certain crustal elements through the atmosphere
Element V Hg Pb Cu Zn Ag Fe Al
155 0.12 32 50 100 2.3 55,000 96,000
0.52 0.009 5 14 51 0.01 200 810
7 0.5 4.1 9 10 0.01 8,800 4,500
5 0.6 38 27 147 1.2 830 2,125
Fossil fuel combustion
Ratio anthropogenic: natural
32 43 88 0.44 641 397
100 to 287 3.4 85 4 5 0.05 4,200 5,900
0.59 to 1.71 1.52 1.48 0.47 0.30 0.14 0.07 0.06
lower value, 1,625 × 109 g V/y, derived from the transport of total suspended sediment, 12.6 × 1015 g/y (25), and the mean concentration of 129 mg V/kg in sediments (26). With respect to chemical weathering, Shiller and Mao (12) observed that world rivers are characterized by relatively constant dissolved V/Si ratio (∼66 × 10−6 mol V/mol Si), suggesting that dissolved V and Si are dominantly derived from chemical weathering of silicate rocks (27). In a detailed study of global weathering, Gaillardet et al. (28) concluded that virtually all dissolved silica in rivers derives from silicate weathering. Following the approach of Shiller and Mao (12), we calculate the V flux from chemical weathering as ∼21 × 109 g/y, using their V/Si correlation and a recent estimate of global dissolved silicate flux (380 × 1012 g/y; ref. 29). Our calculated value for chemical weathering of V is in the middle of the range of earlier estimates for the transport of dissolved V to the oceans in rivers, 8 × 109 g/y to 32 × 109 g/y (12, 30–32). Anthropogenic Sources from Earth’s Crust. We estimate the global anthropogenic V flux from the mining and combustion of coals, exploitation and utilization of petroleum and its products, and direct extraction and processing of V compounds. In general, the exploitation of these resources has increased significantly during the last few decades, which has contributed to a greater mobilization of V through time. Worldwide, ∼85% of V is used as an additive in steel alloys; V is also used as a catalyst in the chemical industry and, increasingly, in V redox batteries (33). Coal production. Global coal production has doubled, from 3,800 million tons/y in the early 1980s to about 8,000 million tons/y in 2014, although, during 2015 and 2016, it decreased to ∼7,300 million tons/y (34). In 2016, China accounted for ∼50% of the global coal production (3,550 million tons/y), followed by India (922 million tons/y) and the United States (661 million tons/y), while all other countries combined accounted for less than 30% of total global production (35). Measured V concentrations in coals worldwide vary from 7 mg/kg to 100 mg/kg, with an estimated global mean of 25 mg/kg (36–39). Archival data on coals from the United States also show a mean V value of 25 mg/kg (median = 20 mg/kg; n = 6,593; ref. 40). The 25-fold increase in the V concentration of coal, compared with the plant tissue from which it forms, likely results from the affinity of V for organic material under reducing conditions and its retention as these materials are compacted and lithified to form coal (e.g., ref. 41). Using this average V concentration in coal, and estimates of global increases in coal production, we estimate that V extraction associated with coal mining has increased from ∼100 × 109 g V/y at the end of the 20th century to 180 × 109 g V/y to 190 × 109 g V/y in 2014–2016 (Fig. 1). During coal combustion, V volatilization is followed by its rapid condensation from flue gas and retention by high-surfacearea fly ash particles (42, 43). V is almost completely accumuSchlesinger et al.
lated on fly ash and other coal combustion residues (CCRs), with little or no V loss due to volatilization (e.g., ref. 44). The V composition in CCRs reflects the V content of the parent coal. The ratio between the volume of CCRs and the volume of utilized coal is about 0.10 (45). Combined with efficient uptake of V in CCRs, this suggests an approximate 10-fold V enrichment in the residual ash compared with coal, averaging ∼250 mg/kg based on the global average V in coal. This estimate is within the range of values (150 mg/kg to 700 mg/kg) reported for V in CCRs (e.g., refs. 37 and 46). In recent decades, installation of high-efficiency cold-side electrostatic precipitators, fabric filters, and wet flue gas desulphurization systems has reduced particulate matter (PM) emissions, especially from coal-fired plants. It has been shown that such “scrubber” systems can eliminate ∼99% of particulate emissions (42). In the United States, most coal-fired plants installed these systems during the 1980s and 1990s (47), while, in China, installation of scrubbers in major power plants increased from almost zero in the early 2000s to about 80% of coal plants in 2010 (48). The dramatic rise in worldwide coal production since the early 2000s due to the accelerated coal mining in China and India suggests higher rates of V mobilization. On the other hand, the transition to modern coal plant technologies with more efficient capturing of PM would restrict the atmospheric emission of V, although the actual effectiveness is still unknown in countries like India or the industrial sector in China. For comparison, atmosphere emissions of Cu and Zn from coal combustion in China increased 1.9% and 1.33% per year, respectively, from 1995 to 2014 (49). In the United States, about half of CCR is used by the cement industry and other beneficial uses (50), and the other half is disposed of in landfills and surface impoundments. Similarly, a large fraction of CCR in China is used by the cement industry (45). The effluent discharged from coal ash ponds has a wide range of V concentrations; a study of effluent in North Carolina showed values ranging from 0.3 μg/L to 53.3 μg/L with a mean value of 10 μg/L (51). Thus, a fraction of the V that is accumulated in CCRs is potentially mobilized to aquatic systems and released to the environment. In North Carolina, 857 million cubic meters of effluent was discharged through regulated outfalls during 2012. Assuming an average V concentration of 10 μg/L in effluent waters, we estimate the annual flux of V from coal ash ponds in North Carolina is 8.6 × 106 g V/y. Since the capacity of coal-powered units in North Carolina (10,430 MW) is equivalent to 3% of the US total capacity (338.7 GW), we estimate the overall V flux from coal ash effluents is about 0.28 × 109 g V/y for the United States, assuming similar effluent discharge practices throughout the country. Since US coal production is about 10% of the global coal production, we estimate that the global V flux to the hydrosphere through effluent release is about 2.8 × 109 g V/y. Recent studies have indicated PNAS | Published online December 11, 2017 | E11093
EARTH, ATMOSPHERIC, AND PLANETARY SCIENCES
From Rauch and Pacyna (88), except for V (this paper) and Hg (107, 108). All data in 109 grams per year.
Fig. 1. The global biogeochemical cycle of V as estimated for current conditions. All values are given in 109 grams per year and rounded to two significant figures from calculations throughout the text.
that coal ash ponds in the United States, and presumably elsewhere, are leaking (52), and thus this value may underestimate the full magnitude of V flux to the hydrosphere from disposal of CCRs in coal ponds and landfills (53). Petroleum production. V and nickel are the two most abundant metals in petroleum, with concentrations up to 1,600 mg/kg and 340 mg/kg, respectively (54). In bitumen, the V level can reach 4,760 mg/kg (55). The V and Ni are associated with organic matter derived from algae that encounters suboxic conditions early in its depositional history (55). The transformation of organic molecules such as plant chlorophylls into organic metallic complexes, known as geoporphyrins, and chelation of Ni2+ and VO2+ to generate metalloporphyrins, preserves V compounds during petroleum diagenesis and catagenesis (56). The differential incorporation of Ni and V into geoporphyrins results in different V and Ni abundances and V/Ni ratios in crude oils. Petroleum from marine environments (defined as Classes A and B) with high sulfur contents is known to contain higher V contents and V/Ni ratios relative to petroleum from mixed and terrestrial settings (Classes C, D, E) (54). V bonds to high molecular weight organic compounds (57), resulting in a positive correlation between V concentration and crude oil density. Consequently, heavy oils are enriched in V relative to light oils. A common unit to express oil density is the American Petroleum Institute gravity (API gravity), in which values above 10 reflect oils with densities lower than water, while oils less than 10 are heavier than water. Data from the US Geological Survey (58) shows greater V in oils with lower API gravity values and differences between conventional oil (average V =16 mg/kg; API gravity = 38), medium oil (V = 98 mg/kg; API gravity = 22), heavy oil (V = 177 mg/kg; API gravity = 16.3), and bitumen (V = 334 mg/kg; API gravity = 5). Similarly, crude oils from Venezuela show an inverse correlation between V concentration and API E11094 | www.pnas.org/cgi/doi/10.1073/pnas.1715500114
gravity, with V concentrations up to 600 mg/kg in biodegradable oils (API gravity < 15), and up to 1,200 mg/kg in heavy (API gravity ∼10) immature oils (59). Consequently, the mobilization of V by petroleum extraction depends strongly on the relative proportions of light and heavy oils. Conventional oil production of light oils, which characterized much of the oil production during the 20th century, generates relatively low V fluxes. Due to increasing worldwide energy demands and technological developments, there has been a significant increase in the exploitation of heavy oils (e.g., Alaskan and Venezuelan oils) and unconventional petroleum resources (e.g., bitumen extraction from tar sands), resulting in significantly greater V mobilization. For example, the production of crude bitumen from Alberta tar sands has more than doubled since 2003, reaching an annual production of 841 million barrels in 2014 (60). A first approximation of the current global V flux from the extraction of petroleum can be obtained using an estimate of the average quality of crude oil (API = 33.6; ref. 61) and an empirical correlation between API gravity and V concentration described by Barbooti (62). These assumptions result in an average V concentration of 28.9 mg/kg in crude oil and a global flux of V from petroleum production of 138 × 109 g/y, assuming 2015 world crude oil production of 34 billion barrels (34, 63). This calculated average V concentration in crude oil does not seem to reflect the increasing worldwide emphasis on heavier grades of crude oil, and the empirical correlation between API gravity and V concentration is based on a limited number of samples (62). Furthermore, V has been shown to be difficult to accurately measure in crude oils (62). As a second approximation, we used recent estimates of the types of petroleum exploited in 2015, finding that about 31% was light sweet crude (averaging 16.2 mg/kg V), 48% was light sour (averaging 98.3 mg/kg V), 13% was heavy sour (averaging Schlesinger et al.
Sinks for V Mobilized from Earth’s Crust Like other low-volatility elements, most V mobilized on land is transported to the oceans, where it is removed by a variety of processes. Removal of sedimentary V in subduction zones can account for 150 × 109 g V/y, calculated from estimates of the mass of subducted sediment (1.3 × 1015 g/y) multiplied by its average V concentration (116 mg/kg; refs. 69 and 70). For comparison, Hope (24) estimates 640 × 109 g of V is removed by subduction of ocean sediment each year. Both estimates are substantially lower than our estimate of the delivery of V to the oceans in suspended sediments (1,625 × 109 g/y; Mobilization of V from Earth’s Crust). The vast majority of suspended sediment delivered to the oceans is not subducted, however, but is instead deposited on passive continental margins and other areas of the ocean floor not currently subject to subduction (71). We estimate the mass of this sequestered ocean sediment as 11.3 × 1015 g/y, following the approach of Rae and Ruff (71) and substituting the recent estimates of sediment supply and subduction noted above (12.6 × 1015 g/y and 1.3 × 1015 g/y, respectively). Multiplying the mass of sediment sequestered by the mean concentration of V in sediment (129 mg V/kg; ref. 26), we calculate that about 1,460 × 109 g V/y is sequestered in oceanic sediment. V is widely recognized for its retention on organic materials, especially under reducing conditions (72, 73). Assuming a concentration of 100 mg V/kg in organic deposits (16), the sequestration of V by the burial of organic matter (of 50% C) in the oceans (0.157 × 1012 g/y; ref. 74) is only about 0.03 × 109 g V/y. Thus, despite its affinity for organic matter, V is largely deposited in inorganic sediments in the marine environment. Oxic sediments Schlesinger et al.
Exchange of V with Earth’s Atmosphere Natural Sources in the Atmosphere. V is delivered to the atmosphere as aerosols derived from various natural processes, each of which represents a small source relative to rock weathering, as discussed above. Current evidence suggests that the largest source of atmospheric V aerosols is wind erosion, with previous estimates ranging from 16 × 109 g V/y to 27 × 109 g V/y (15, 81). To this range, we provide a larger estimate of 155 × 109 g V/y, based on the eolian mobilization of crustal materials with particle sizes of <10 μm (i.e., PM10; 1,600 × 1012 g/y; ref. 82) multiplied by the mean concentration of V in Earth’s crust (97 mg/kg; refs. 1 and 11). The next largest source of atmospheric V is subaerial volcanic emissions. A common approach to calculate the global flux of trace metals from volcanic activity has been to estimate their metal/sulfur ratios and multiply by the relatively well-investigated volcanic sulfur emission. Previous estimates using this method suggest a global range for V emissions of 0.1 × 109 g/y to 5.5 × 109 g/y (ref. 15, adjusted for an error noted by ref. 83). While this approach and refinements upon it are logical for estimating the volcanic flux of volatile trace metals (e.g., Hg, Pb, Cd; ref. 84), it is likely less appropriate for estimating fluxes of lowvolatility metals, such as V, which are emitted primarily in volcanic ash (e.g., refs. 78, 83, and 85–87). Therefore, building on the approach of Duce and Hoffman (81), we calculate V/Al ratios of volcanic particulates (using a range of upper crustal and andesitic values of 15.3 to 17 wt.% Al2O3; and 97 mg/kg V to 175 mg/kg V; ref. 1) and multiply by two estimates of global volcanic Al emissions [4,500 × 109 g/y (88); 13,280 × 109 g/y (83)]. Using the lower Al emission value, we calculate global volcanic V fluxes ranging from 5 × 109 g V/y to 10 × 109 g V/y, similar to the estimate of Duce and Hoffman (81), while the higher Al value suggests a range of 14 × 109 g V/y to 29 × 109 g V/y. Wildfires are also a source of V in the atmosphere, estimated between 1.8 × 109 g/y and 13 × 109 g/y (15, 24). We calculate wildfire emissions of V by assuming that 5 × 1015 g/y of biomass are burned annually (89), volatilizing V from tissues that have a mean PNAS | Published online December 11, 2017 | E11095
tend to release V to seawater, while suboxic (including buried) sediments adsorb and retain dissolved V (75). The mean concentration of V dissolved in seawater, 1.8 μg/L, is higher than that in river waters (0.7 μg/L), implying additional sources of V in seawater. Some V is likely released from suspended or accumulating sediments as they experience the higher pH of seawater and from sedimentary environments with oxic conditions (72, 76). In contrast, hydrothermal vent systems appear to be a net sink rather than a source for V to seawater, with hydrothermal sediment accumulations ranging from 7.1 × 109 g V/y (75, 77) to 22 × 109 g V/y to 28 × 109 g V/y (32, 75). Further, while undersea volcanic eruptions at midocean ridges and oceanic volcanoes can influence the seawater composition of various volatile trace elements, V mobilized during such eruptions appears to be rapidly deposited on surrounding seafloor (78, 79). The sum of the dissolved V in rivers (∼21 × 109 g V/y to 30 × 109 g V/y) and the release from oxic sediments (∼15 × 109 g V/y to 41 × 109 g V/y; ref. 75) can accommodate the estimated uptake by hydrothermal deposits (22 109 g V/y to 28 × 109 g V/y) and suboxic marine sediments (22 × 109 g V/y; ref. 75), so that the budget for dissolved V in the ocean is reasonably well balanced. The mean residence time for dissolved V in the oceans is about 130,000 y, with respect to river inputs [i.e., 2.7 × 1015 g of V dissolved in seawater (13) divided by 21 × 109 g dissolved V in rivers (this paper)]. The estimated mean residence time decreases to about 42,000 y if the influx includes rivers, atmospheric deposition, and release from oxic sediments (Fig. 1). For comparison, Whitfield and Turner (80) estimate a mean residence time of about 100,000 y, based on the partitioning of V between seawater and marine sediments.
EARTH, ATMOSPHERIC, AND PLANETARY SCIENCES
177 mg/kg V), and 4% was natural bitumen (averaging 334 mg/kg V) (34, 58, 63). From these estimates, we calculate a current mobilization of about 410 × 109 g V/y by extraction of petroleum from Earth’s crust, a value about threefold that estimated above. Using this second approach, we also estimate that this flux has approximately doubled over the past two decades, roughly consistent with the annual flux (165 × 109 g V/y) calculated by Monakhov et al. (64) for the year 2000. The increasing V flux over this time is due to both the 1.5-fold increase in global oil production as well as the increasing proportion of high-V petroleum fuels in the mix. Together, these two approximations constrain the anthropogenic flux of V from petroleum extraction to lie between 138 × 109 g/y and 410 × 109 g/y. While V is enriched in crude oils, limited data suggest that the V content in produced water generated during the production of oil is negligible. For example, oil-produced water from the Gulf of Mexico contains V up to 1 μg/L, and most water samples had values below detection limit (65). Consequently, unlike other elements that are enriched in the formation water associated with oil, and thus would contribute to anthropogenic fluxes, such as boron (66), cadmium (Cd), and lead (Pb) (65), it seems that the V flux through produced water is very small. Mining and formation of V chemicals. V is produced by mining of vanadiferous titanomagnetite ores, and extracted as a byproduct in the processing of iron ore, phosphates, and uranium. Data presented by Monakhov et al. (64) suggest that the purposeful mining of V-bearing ores mobilized about 130 × 109 g V/y in 2000, with about 43% from virgin V ores and 56% produced as a byproduct. Analyses by the US Geological Survey show that world V production, primarily from mining, has more than doubled over the past two decades (from 34 × 109 g V/y in 1994 to 82 × 109 g V/y in 2014), with China, South Africa, and Russia as the major producers (67, 68). We suggest that the total mobilization of V from Earth’s crust by human activities is currently (2015–2016) 730 × 109 g/y, with 18% from direct and byproduct mining activities and, as described above, 56% from petroleum production, and 26% from coal.
concentration of 1 mg/kg (10), and thus delivering 5 × 109 g V/y to the atmosphere. Finally, while sea salt aerosols have been shown to be a significant atmospheric source for some elements (e.g., Na, B, Cl), little V is mobilized by this process. We calculate that production of sea salt aerosols at the ocean’s surface (10.1 × 1015 g/y; ref. 82) multiplied by the ratio of V to total dissolved solids in seawater (5.14 × 10−8) would inject 0.52 × 109 g V/y into the marine atmosphere—a trivial flux in its global cycle. Zoller et al. (14) indicate an even smaller value, 0.012 × 109 g V/y. Anthropogenic Additions of V to Atmosphere. Coal combustion. To provide a first estimate of the atmospheric
emission of V from coal combustion, we use US Environmental Protection Agency data (90) that quantifies the masses PM10 and PM2.5 that are annually emitted from electric utilities in the United States, combined with our estimate of the average V concentration in CCRs (250 mg/kg). Our calculations indicate that, despite the doubling in national coal production in the United States from 500 million tons/y to 600 million tons/y during the early 1970s to ∼1,000 million tons/y in the mid-2000s, the atmospheric flux of V as determined by the PM mass flux declined from 0.4 × 109 g V/y in the early 1970s to 0.1 × 109 g V/y in the late 2000s and recent years, representing a fourfold reduction. This trend reflects improvements in the use and efficiency of ESP and other scrubber systems in reducing PM emissions to the atmosphere. Since global PM emission from coal plants is unknown, we use the US dataset to approximate the global flux under two different scenarios. In the first scenario, we assume that the reduction of PM emission that was achieved in the US is valid for the rest of the world; in the second scenario we assume that the conditions that characterize high PM emission in the US in the early 1970s apply to the rest of the world today. In recent years, US coal production represents about 10% of global coal production and thus we can assume that V fugitive emissions in the US are also equivalent to 10% of the global V emission. Under Scenario 1 (global reduction of PM emission), the global V atmospheric flux from coal combustion is about 1.1 × 109 g V/y (i.e., 10 times the current US emission). According to Scenario 2 (global emission mimics the US emission during early 1970s), the global V atmospheric flux from coal combustion is about 4.4 × 109 g V/y (i.e., 10 times the US emission during early 1970s normalized to the current global coal production). Petroleum Sources of V to the Atmosphere. Monakhov et al. (64) estimated that the annual V flux to the atmosphere from petroleum combustion during the year 2000 was 113 × 109 g/y, which represented about 69% of the V contained in petroleum extracted from the ground (165 × 109 g V/y). Since then, global oil production has increased, and the V content in extracted petroleum may lie between 138 × 109 g V/y and 410 × 109 g V/y (see above). While improvements have been made in developed countries to capture a greater proportion of the particulate flux from all forms of fossil fuel combustion (e.g., ref. 91), this is offset by increased emissions in developing countries and the absence of emissions regulations for various types of fuel in some developed countries. Visschedijk et al. (92) report that 69% of the emissions of V are not subject to scrubbing in Europe. If we assume that 69% of the V content from petroleum combustion is emitted to the atmosphere, we calculate that the current flux may be as high as 283 × 109 g V/y. This flux now dominates the emission of V to the atmosphere. The anthropogenic additions of V to the atmosphere may now exceed natural sources by a factor of 1.7 (Table 1)—larger than the similar ratios for Hg and Pb, but somewhat lower than estimated by previous workers (24, 93). E11096 | www.pnas.org/cgi/doi/10.1073/pnas.1715500114
An increasing fraction of V atmospheric emissions estimated above derives from combustion of unconventional petroleum products, such as residual/heavy oils and petroleum coke. The selective incorporation of V into high molecular weight organic compounds (57) controls the preservation and accumulation of V in residual organic materials during distillation and refining. As a result, V is enriched in the residues during the conversion and refinement of oil and bitumen. Global production of residual/heavy oil was about 467 million metric tons per year in 2014, although, in recent years, demand has decreased in most sectors (94). Other notable residue products include petroleum coke, a solid carbon-rich material increasingly used as a combustion fuel due to its high carbon content, low cost, and growing abundance (95). V enrichment in these residues is often marked; for example, petroleum coke produced during distillation of bitumen from the Athabasca Oil Sands Region in Canada shows a mean V concentration of 1,280 ± 120 mg/kg (96), while bitumen from this region ranges from 60 mg/kg to 640 mg/kg (97). The global production of petroleum coke rose to about 124 million metric tons per year in 2014, with the United States producing about 47% (94). Approximately 70% of worldwide production is used as a source of combustible fuel (94). In some cases, petroleum coke substitutes for coal; for example, in 2002, about 10% of power plants in Kentucky utilized petroleum coke instead of coal (98). Fly ash originating from combustion of petroleum coke is highly enriched in V, with reported concentrations between 2,000 mg/kg and 40,000 mg/kg (99, 100). Consequently, fugitive emission of fly ash particles from petroleum coke combustion would result in high atmospheric emission of V. In the absence of effective scrubber systems, combustion of heavy oils and petroleum coke can dominate regional atmospheric V emissions. For example, a study of 2005 emissions in northwestern Europe showed that combustion of residual oil (e.g., ships and petroleum refineries) and petroleum coke (power plants and refineries) contributed 80% and 18%, respectively, of the total atmospheric emission of V in that region (1.57 × 109 g V/y; ref. 92). Studies in Spain (101) and in the United States (14, 102) show similar findings. Based on estimates of the current mix of petroleum products exploited (described above), we estimate that unconventional fuels, with their higher V concentrations, represent about 40.5% (166 × 109 g V/y) of the total V mobilized to the atmosphere from the extraction of petroleum (410 × 109 g V/y). If we also assume that, of the V emitted to the atmosphere from burning petroleum products (283 × 109 g V/y), 40.5% originates from unconventional fuels, we calculate that 115 × 109 g V/y are released to the atmosphere from the combustion of residual/heavy oil, petroleum coke, and other unconventional fuels. Sinks of V from the Atmosphere Measurements of the concentration of V in rainfall collected in remote areas and in continental areas subject to human activities range over four orders of magnitude (Table 2). Whenever possible, we provide the ratio of V to Al in these samples to provide an estimate of the “enrichment factor” of V in rainwater versus the V in Earth’s crust. Most of the data from remote areas have enrichment factors of <10, whereas those from areas of human activities have factors ranging up to 1,000. For remote sites, the median concentration of V in rainfall, 0.03 μg/L and global precipitation of 5 × 1017 L/y yields an estimate of global atmospheric deposition in rainfall of ∼15 × 109 g V/y in preindustrial times. This atmospheric deposition of V might have accounted for a small portion of the dissolved V in rivers. For modern conditions, we assume that the wet deposition over the oceans has not been affected by human activities and thus could be 11 × 109 g V/y. For the land, the median of all concentration values for V in modern precipitation, ∼0.4 μg/L, would imply deposition of 45 × 109 g V/y. The median of samples with enrichment factors of <10 would imply deposition of 35 × 109 g V/y from the atmosphere and a Schlesinger et al.
Table 2. Concentrations of V in precipitation collected in various locations
Remote areas East Antarctica Coats Land, East Antarctica East Antarctica Greenland Enewetak Atoll Tibetan Plateau, China Bermuda Hawai’i (volcano quiescent) Korea (rural) Alps, Switzerland Hong Kong, China Alps, Switzerland Continental areas with moderate Hiroshima, Japan Newark, New Jersey, USA Pensacola, Florida Lamto, Ivory Coast Lhasa, Tibet, China Japan (four sites) Tibetan Plateau, China Culpepper, Virginia Reston, Virginia Tokyo, Japan Delaware Montreal, Canada Oleiros A Coruna, Spain South China Massachusetts, Rhode Island Northern China Ankara, Turkey Kollo, Niger Southern Jordan Singapore, Malaysia Nanjing, China Mexico City, Mexico
Ref. 109 110 105 Herron et al. 1975 as cited by ref. 81 111 112 113 85 114 115 116 115 117 118 119 120 121 122 123 124 125 126 113 127 128 129 130 131 Li et al. 2012 as cited by ref. 121 132 120 133 134 Tang 2007 as cited by ref. 121 135
*Versus Al; namely, EF = (V/Al)rain/(V/Al)crust.
substantial enhancement (∼9.0×) due to human activities. The latter amounts to slightly more than 20% of estimated gross sources of V in the atmosphere, suggesting that a substantial amount of dry deposition must occur near sources of human emissions (7, 103, 104). However, some V is carried long distances in the atmosphere. For example, dissolved and particulate phases are roughly equally represented in Antarctic snow (105), and V in Greenland snow shows a strong relation to crustal dusts, which are the largest natural source of V in the atmosphere (106). In the remote atmosphere, concentrations of V are typically about 1 × 10−9 g/m3, which would result in a global dry deposition of 16 × 109 g/y, assuming a nominal dry deposition velocity of 0.1 cm/s for small particles. Our calculation of the sum of wet and dry deposition (72 × 109 g/y) is less than the estimated sources to the atmosphere, leaving us with an unbalanced budget for V in the atmosphere. In urban areas, airborne concentrations can exceed 10 × 10−9 g/m3 (14, 92, 101, 102), so the atmospheric budget may be balanced by dry deposition near emission sources. Conclusions Overall, our data provide insights on the current anthropogenic V fluxes from coal mining and combustion; crude oil exploration, processing, and combustion; and ore mining and industrial processing. Since the beginning of the 21st century, the increasing Schlesinger et al.
use of V-rich heavy oil, bitumen, and petroleum coke from unconventional petroleum resources and production residues has changed the V fluxes, such that the V derived from petroleum products is now the largest global flux (56%), followed by coal (26%) and V mining (18%). Data on atmospheric V emissions are limited, but the accumulation of V in heavy oils and bitumen residuals, such as petroleum coke, poses a risk associated fugitive emission of V-rich PM, with unknown impacts on human health. Fig. 1 summarizes the salient points for the global biogeochemical cycle of V: i) Humans have potentially increased the total mobilization of V from Earth’s crust by a factor of >1.45, as a result of mining of V ores and the extraction and combustion of fossil fuels. The dissolved flux to the seas has likely increased by ∼15%. ii) Human emissions of V to the atmosphere now exceed background levels by as much as a factor of 1.7, and we are likely to have increased the deposition of V from the atmosphere by a similar amount since preindustrial times. Much of this is probably derived from airborne emissions during the combustion of fossil fuels, and this flux is likely to increase in the near future as humans pursue sources of petroleum from heavy oils and bitumen. iii) The human impact on the atmospheric flux of V may exceed that for Hg and Pb (Table 1), primarily due to the significant PNAS | Published online December 11, 2017 | E11097
EARTH, ATMOSPHERIC, AND PLANETARY SCIENCES
inputs and outputs, and a mean residence time for dissolved V in seawater of about 130,000 y with respect to inputs from rivers.
enrichment of V in heavy crude oil and byproducts. While future air pollution control could mitigate atmospheric V emission from combusted fossil fuels, the expected increasing utilization of unconventional petroleum resources could further exacerbate the anthropogenic V flux in the future. iv) The budget of dissolved V in the oceans is remarkably well balanced—with about 40 × 109 g V/y to 50 × 109 g V/y
ACKNOWLEDGMENTS. We thank Terry Plank for her help with the calculation for the subduction of V in marine sediments, Robert Duce for his insights on dry deposition flux, and Stan Coffman for his drawing that summarizes the global V cycle.
1. Rudnick RL, Gao S (2014) Composition of the continental crust. Treatise on Geochemistry, eds Turekian KK, Holland HD (Elsevier, Amsterdam), 2nd Ed, pp 1–51. 2. Bellenger JP, et al. (2014) Possible contributions of alternative nitrogenase to nitrogen fixation by asymbiotic N2-fixing bacteria in soils. Soil Biol Biochem 69: 413–420. 3. Darnajoux R, et al. (2017) Biological nitrogen fixation by alternative nitrogenases in boreal cyanolichens: Importance of molybdenum availability and implications for current biological nitrogen fixation estimates. New Phytol 213:680–689. 4. Zhang X, et al. (2016) Alternative nitrogenase activity in the environment and nitrogen cycle implications. Biogeochem Lett 127:189–198. 5. Wever R, Tromp MGM, Krenn BE, Marjani A, Vantol M (1991) Brominating activity of the seaweed, Ascophyllum nodosum–Impact on the biosphere. Environ Sci Technol 25:446–449. 6. Johnson TL, Brahamsha B, Palenik B, Muehle J (2015) Halomethane production by vanadium-dependent bromoperoxidase in marine Synechococcus. Limnol Oceanogr 60:1823–1835. 7. Imtiaz M, et al. (2015) Vanadium, recent advancements and research prospects: A review. Environ Int 80:79–88. 8. Nielsen FH, Uthus EO (1990) III. The essentiality and metabolism of vanadium. Vanadium in Biological Systems: Physiology and Biochemistry, ed Chasteen ND (Kluwer, Amsterdam), pp 51–62. 9. Crans DC, Smee JJ, Gaidamauskas E, Yang L (2004) The chemistry and biochemistry of vanadium and the biological activities exerted by vanadium compounds. Chem Rev 104:849–902. 10. Bowen HJM (1966) Trace Elements in Biochemistry (Academic, New York). 11. Wedepohl KH (1995) The composition of the continental crust. Geochim Cosmochim Acta 59:1217–1232. 12. Shiller AM, Mao LJ (2000) Dissolved vanadium in rivers: Effects of silicate weathering. Chem Geol 165:13–22. 13. Huang J-H, Huang F, Evans L, Glasauer S (2015) Vanadium: Global (bio)geochemistry. Chem Geol 417:68–89. 14. Zoller WH, Gordon GE, Gladney ES, Jones AG (1973) Sources and distribution of vanadium in the atmosphere. Trace Elements in the Environment, ed Kothny EL (Am Chem Soc, Washington, DC), pp 31–47. 15. Nriagu JO (1989) A global assessment of natural sources of atmospheric trace metals. Nature 338:47–49. 16. Groen JC, Craig JR (1994) The inorganic geochemistry of coal, petroleum, and their gasification/combustion products. Fuel Process Technol 40:15–48. 17. Zuliani JE, et al. (2016) Characterization of vanadium in oil sands fluid petroleum coke using electron microscopy. Fuel 178:124–128. 18. Gummow B (2011) Vanadium: Environmental pollution and health effects. Encyclopedia of Environmental Health, ed Nriagu JO (Elsevier, Amsterdam), pp 628–636. 19. Schiffer S, Liber K (2017) Toxicity of aqueous vanadium to zooplankton and phytoplankton species of relevance to the athabasca oil sands region. Ecotoxicol Environ Saf 137:1–11. 20. Cloy JM, Farmer JG, Graham MC, MacKenzie AB (2011) Scottish peat bog records of atmospheric vanadium deposition over the past 150 years: Comparison with other records and emission trends. J Environ Monit 13:58–65. 21. Allan M, et al. (2013) High-resolution reconstruction of atmospheric deposition of trace metals and metalloids since AD 1400 recorded by ombrotrophic peat cores in Hautes-Fagnes, Belgium. Environ Pollut 178:381–394. 22. US Geological Survey (2015) Mineral commodity summaries—Vanadium. Available at https://minerals.usgs.gov/minerals/pubs/commodity/vanadium/mcs-2016-vanad.pdf. Accessed July 8, 2016. 23. Goonan TG (2011) Vanadium recycling in the United States in 2004. Flow Studies for Recycling Metal Commodities in the United States, US Geological Survey Circular 1196, ed Sibley SF (US Geol Survey, Washington, DC), pp S1–S17. 24. Hope BK (2008) A dynamic model for the global cycling of anthropogenic vanadium. Global Biogeochem Cycles 22:GB4021. 25. Syvitski JPM, Vörösmarty CJ, Kettner AJ, Green P (2005) Impact of humans on the flux of terrestrial sediment to the global coastal ocean. Science 308:376–380. 26. Viers J, Dupré B, Gaillardet J (2009) Chemical composition of suspended sediments in world rivers: New insights from a new database. Sci Total Environ 407:853–868. 27. Gardner CB, et al. (2017) Molybdenum, vanadium, and uranium weathering in small mountainous rivers and rivers draining high-standing islands. Geochim Cosmochim Acta 219:22–43. 28. Gaillardet J, Depré B, Louvat P, Allègre CJ (1999) Global silicate weathering and CO2 consumption rates deduced from the chemistry of large rivers. Chem Geol 159:3–30. 29. Beusen AHW, Bouwman AF, Dürr HH, Dekkers AL, Hartmann J (2009) Global patterns of dissolved silica export to the coastal zone: Results from a spatially explicit global model. Global Biogeochem Cycles 23:GB0A02.
30. Bertine KK, Goldberg ED (1971) Fossil fuel combustion and the major sedimentary cycle. Science 173:233–235. 31. Shiller AM, Boyle EA (1987) Dissolved vanadium in rivers and estuaries. Earth Planet Sci Lett 86:214–224. 32. Elderfield H, Schultz A (1996) Mid-ocean ridge hydrothermal fluxes and the chemical composition of the ocean. Annu Rev Earth Planet Sci 24:191–224. 33. Moskalyk RR, Alfantazi AM (2003) Processing of vanadium: A review. Miner Eng 16: 793–805. 34. British Petroleum (2017) British Petroleum statistical review of the world energy. Available at https://www.bp.com/en/global/corporate/energy-economics/statisticalreview-of-world-energy.html. Accessed July 4, 2017. 35. Enerdata (2017) Global energy statistical yearbook. Available at https://yearbook. enerdata.net/. Accessed July 4, 2017. 36. Dai S, et al. (2012) Geochemistry of trace elements in Chinese coals: A review of abundances, genetic types, impacts on human health, and industrial utilization. Int J Coal Geol 94:3–21. 37. Ketris MP, Yudovich YE (2009) Estimations of Clarkes for Carbonaceous biolithes: World averages for trace element contents in black shales and coals. Int J Coal Geol 78:135–148. 38. Wang J, et al. (2008) Statistical analysis of the concentrations of trace elements in a wide diversity of coals and its implications for understanding elemental modes of occurrence. Fuel 87:2211–2222. 39. Xu M, et al. (2004) Status of trace element emission in a coal combustion process: A review. Fuel Process Technol 85:215–237. 40. US Geological Survey (2015) The U.S. Geological Survey coal quality (COALQUAL) database, version 3.0: Data Series 975. Available at pubs.er.usgs.gov/publication/ ds975. Accessed July 4, 2017. 41. Scott C, Stack JF, Kelley KD (2017) The hyper-enrichment of V and Zn in black shales of the Late Devonian-Early Mississippian Bakken formation (USA). Chem Geol 452: 24–33. 42. Meij R, Winkel BH (2009) Trace elements in world steam coal and their behavior in Dutch coal-fired power stations: A review. Int J Coal Geol 77:289–293. 43. Sahoo PK, Kim K, Powell MA, Equeenuddin SM (2016) Recovery of metals and other beneficial products from coal fly ash: A sustainable approach for fly ash management. Intern J Coal Sci Technol 3:267–283. 44. Wang J, Tomita A (2003) A chemistry on the volatility of some trace elements during coal combustion and pyrolysis. Energy Fuels 17:954–960. 45. Heidrich C, Feuerborn H-J, Weir A (2013) Coal Combustion Products: A Global Perspective, World of Coal Ash Conference (Am Coal Ash Assoc, Farmington Hills, MI). Available at energy.caer.uky.edu/AshSymposium/AshLibraryAgenda.asp. Accessed August 26, 2017. 46. Tiwari M, Sahu SK, Bhangare RC, Ajmal PY, Pandit GG (2014) Elemental characterization of coal, fly ash, and bottom ash using an energy dispersive X-ray fluorescence technique. Appl Radiat Isot 90:53–57. 47. Environmental Health & Engineering (2011) Emissions of hazardous air pollutants from coal-fired power plant: EH&E Report 17505. Available at https://www.csu. edu/cerc/researchreports/documents/EmissionsOfHazardousAirPollutantsFromCoalFiredPowerPlants2011.pdf. Accessed July 4, 2017. 48. Wang S, Hao J (2012) Air quality management in China: Issues, challenges, and options. J Environ Sci (China) 24:2–13. 49. Li R, et al. (2017) Atmospheric emissions of Cu and Zn from coal combustion in China: Spatio-temporal distribution, human health effects, and short-term prediction. Environ Pollut 229:724–734. 50. US Environmental Protection Agency (2017) Coal combustion residuals. Available at https://www.epa.gov/coalash. Accessed July 4, 2017. 51. Ruhl L, et al. (2012) The impact of coal combustion residue effluent on water resources: A North Carolina example. Environ Sci Technol 46:12226–12233. 52. Harkness JS, Sulkin B, Vengosh A (2016) Evidence for coal ash ponds leaking in the southeastern United States. Environ Sci Technol 50:6583–6592. 53. Izquierdo M, Querol X (2012) Leaching behavior of elements from coal combustion fly ash: An overview. Int J Coal Geol 94:54–66. 54. Barwise AJG (1990) Role of nickel and vanadium in petroleum classification. Energy Fuels 4:647–652. 55. Lewan MD, Mynard JB (1982) Factors controlling enrichment of vanadium and nickel in the bitumen of organic sedimentary rocks. Geochim Cosmochim Acta 46: 2547–2560. 56. Filby RH, Van Berkel GJ (1987) Geochemistry of metal complexes in petroleum, source rocks, and coals: An overview. Metal Complexes in Fossil Fuels, eds Filby RH, Branthaver JF (Am Chem Soc, Washington, DC), pp 2–39. 57. Lewan MD (1984) Factors controlling the proportionality of vanadium to nickel in crude oils. Geochim Cosmochim Acta 48:2231–2238.
PNAS | Published online December 11, 2017 | E11099
91. Cusack M, Alastuey A, Pérez N, Pey J, Querol X (2012) Trends of particulate matter (PM2.5) and chemical composition at a regional background site in the western Mediterranean over the last nine years (2002-2010). Atmos Chem Phys 12: 8341–8357. 92. Visschedijk AHJ, Denier van der Gon HAC, Hulskotte JHJ, Quass U (2013) Anthropogenic vanadium emissions to air and ambient air concentrations in North-West Europe. Available at dx.doi.org/10.1051/e3sconf/20130103004. Accessed July 4, 2017. 93. Nriagu JO, Pirrone N (1998) Emissions of vanadium into the atmosphere. Vanadium in the Environment. Part I: Chemistry and Biochemistry, ed Nriagu JO (Wiley, New York), pp 25–36. 94. International Energy Agency (2016) World Energy Statistics (2016) (OECD Publ, Paris). 95. International Energy Agency (2016) Oil Information (2016) (OECD Publ, Paris). 96. Nesbitt JA, Lindsay MB (2017) Vanadium geochemistry of oil sands fluid petroleum coke. Environ Sci Technol 51:3102–3109. 97. Funk EW, Gomez E (1977) Determination of vanadium in Athabasca bitumen and other heavy hydrocarbons by visible spectrometry. Anal Chem 49:972–974. 98. Hower J, Thomas GA, Mardon SM, Trimbleet AS (2005) Impact of co-combustion of petroleum coke and coal on fly ash quality: Case study of a Western Kentucky power plant. Appl Geochem 20:1309–1319. 99. Henke KR (2005) Trace Element Chemistry of Fly Ashes from Co-Combusted Petroleum Coke and Coals, 2005 World of Coal Ash (WOCA) Conference (Am Coal Ash Assoc, Farmington Hills, MI). Available at www.http://energy.caer.uky.edu/AshSymposium/ AshLibraryAgenda.asp. Accessed August 26, 2017. 100. Lai A (2010) Ferrovanadium production from heavy fuel oil fly ash and BOF: TA Report AES/RE/10-11. Available at https://repository.tudelft.nl/islandora/object/uuid: d09a3aa8-93b4-46e9-af3b-eac8701badf3/datastream/OBJ. Accessed July 4, 2017. 101. Moreno T, et al. (2010) Variations in vanadium, nickel and lanthanoid element concentrations in urban air. Sci Total Environ 408:4569–4579. 102. Peltier RE, Lippmann M (2010) Residual oil combustion: 2. Distributions of airborne nickel and vanadium within New York City. J Expo Sci Environ Epidemiol 20:342–350. 103. Baker AR, Thomas M, Bange HW, Sanchez EP (2016) Soluble trace metals in aerosols over the tropical south-east Pacific offshore Peru. Biogeosciences 13:817–825. 104. Omrani M, Ruban V, Ruban G, Lamprea K (2017) Assessment of atmospheric trace metal deposition in urban environments using direct and indirect measurement methodology and contributions from wet and dry depositions. Atmos Environ 168: 101–114. 105. Grotti M, et al. (2015) Year-round record of dissolved and particulate metals in surface snow at Dome Concordia (East Antarctica). Chemosphere 138:916–923. 106. Lai AM, Shafer MM, Dibb JE, Polashenski CM, Schauer JJ (2017) Elements and inorganic ions as source tracers in recent Greenland snow. Atmos Environ 164: 205–215. 107. Selin NE (2009) Global biogeochemical cycling of mercury: A review. Annu Rev Environ Resour 34:43–63. 108. Sen IS, Peucker-Ehrenbrink B (2012) Anthropogenic disturbance of element cycles at the Earth’s surface. Environ Sci Technol 46:8601–8609. 109. Hur SD, et al. (2007) Seasonal patterns of heavy metal deposition to the snow on Lambert Glacier basin, East Antarctica. Atmos Environ 41:8567–8578. 110. Planchon FAM, et al. (2002) Short-term variations in the occurrence of heavy metals in Antarctic snow from Coats Land since the 1920s. Sci Total Environ 300:129–142. 111. Arimoto R, Duce RA, Ray BJ, Unni CK (1985) Atmospheric trace-elements at Enewetak Atoll. 2. Transport to the ocean by wet and dry deposition. J Geophys Res Atmos 90:2391–2408. 112. Cong Z, Kang S, Zhang Y, Li X (2010) Atmospheric wet deposition of trace elements to central Tibetan Plateau. Appl Geochem 25:1415–1421. 113. Church TM, et al. (1984) The wet deposition of trace metals to the western Atlantic ocean at the mid-Atlantic coast and on Bermuda. Atmos Environ 18:2657–2664. 114. Kim J-E, Han Y-J, Kim PR, Holsen TM (2012) Factors influencing atmospheric wet deposition of trace elements in rural Korea. Atmos Res 116:185–194. 115. Gabrieli J, et al. (2011) Impact of Po valley emissions on the highest glacier of the eastern European Alps. Atmos Chem Phys 11:8087–8102. 116. Tanner PA, Wong AYS (2000) Soluble trace metals and major ionic species in the bulk deposition and atmosphere of Hong Kong. Water Air Soil Pollut 122:261–279. 117. Takeda K, Marumoto K, Minamikawa T, Sakugawa H, Tujiwara K (2000) Three-year determination of trace metals and the lead isotope ratio in rain and snow depositions collected in Higashi-Hiroshima, Japan. Atmos Environ 34:4525–4535. 118. Song F, Gao Y (2009) Chemical characteristics of precipitation at metropolitan Newark in the U.S. East Coast. Atmos Environ 43:4903–4913. 119. Landing WM, Caffrey JM, Nolek SD, Gosnell KJ, Parker WC (2010) Atmospheric wet deposition of mercury and other trace elements in Pensacola, Florida. Atmos Chem Phys 10:4867–4877. 120. Freydier R, Dupre B, Lacaux JP (1998) Precipitation chemistry in intertropical Africa. Atmos Environ 321:749–765. 121. Guo J, et al. (2015) Seasonal variations of trace elements in precipitation at the largest city in Tibet, Lhasa. Atmos Res 153:87–97. 122. Sakata M, Asakura K (2009) Factors contributing to seasonal variations in wet deposition fluxes of trace elements at sites along Japan sea coast. Atmos Environ 43: 3867–3875. 123. Liu B, et al. (2013) Wet precipitation chemistry at a high-altitude site (3,326 m a.s.l.) in the southeastern Tibetan Plateau. Environ Sci Pollut Res Int 20:5013–5027. 124. Engle MA, Kolker A, Mose DE, East JA, McCord JD (2008) Summary of mercury and trace element results in precipitation from the Culpepper, Virginia, Mercury Deposition Network Site (VA-08), 2002-2206 (US Geol Surv, Boulder, CO), US Geol Surv Open-File Rep 2008-1232.
EARTH, ATMOSPHERIC, AND PLANETARY SCIENCES
58. Meyer RF, Attanasi ED, Freeman PA (2007) Heavy oil and natural bitumen resources in geological basins of the world: US Geological Survey Open-File Report 2007-1084. Available at https://pubs.usgs.gov/of/2007/1084/. Accessed July 4, 2017. 59. López L, Mónaco SL (2017) Vanadium, nickel and sulfur in crude oils and source rocks and their relationship with biomarkers: Implications for the origin of crude oils in Venezuelan basins. Org Geochem 104:53–68. 60. Alberta Government (2015) Alberta Official Statistics: Total Bitumen Production. Updated July 27, 2015. Available at www.finance.alberta.ca/aboutalberta/osi/aos/ data/Total-Bitumen-Production-AB.pdf. 61. World Oil Outlook (2015) Organization Petroleum Exporting Countries (World Oil Outlook, Vienna). 62. Barbooti M (2015) Evaluation of analytical procedures in the determination of trace metals in heavy crude oils by flame atomic absorption spectrophotometry. Am J Anal Chem 6:325–333. 63. US Department of Energy, Energy Information Administration (2006) Annual Energy Outlook 2006, with Projections to 2030 (US Dep Energy, Washington, DC). Accessed August 8, 2017. 64. Monakhov IN, Khromov SV, Chernousov PI, Yu S, Yusfin YS (2004) The flow of vanadium-bearing materials in industry. Metallurgist 48:381–385. 65. Trefry JH, Naito KL, Trocine RP, Metz S (1995) Distribution and bioaccumulation of heavy metals from produced water discharges to the Gulf of Mexico. Water Sci Technol 32:31–36. 66. Schlesinger WH, Vengosh A (2016) Global boron cycle in the anthropocene. Global Biogeochem Cycles 30:219–230. 67. Reese RG (1997) Vanadium: U.S. Geological Survey Minerals Information 1997. Available at https://minerals.usgs.gov/minerals/pubs/commodity/vanadium/. Accessed July 24, 2017. 68. Polyak DE (2016) Vanadium: U.S. Geological Survey 2015 Minerals Yearbook. Available at https://minerals.usgs.gov/minerals/pubs/commodity/vanadium/. Accessed July 4, 2017. 69. Plank T, Langmuir CH (1998) The chemical composition of subducting sediment and its consequences for the crust and mantle. Chem Geol 145:325–394. 70. Plank T (2014) The chemical composition of subducting sediments. Treatise on Geochemistry, eds Turekian KK, Holland HD (Elsevier, New York), 2nd Ed, pp 607–629. 71. Rae DK, Ruff LJ (1996) Composition and mass flux of sediment entering the world’s subduction zones: Implications for global sediment budgets, great earthquakes, and volcanism. Earth Planet Sci Lett 140:1–12. 72. Emerson SR, Huested S (1991) Ocean anoxia and the concentrations of molybdenum and vanadium in seawater. Mar Chem 34:177–196. 73. Telfeyan K, et al. (2017) Arsenic, vanadium, iron, and manganese biogeochemistry in a deltaic wetland, southern Louisiana, USA. Mar Chem 192:32–48. 74. Berner RA (1982) The burial of organic carbon and pyrite sulfur in the modern ocean— Its geochemical and environmental significance. Am J Sci 282:451–473. 75. Morford JL, Emerson S (1999) The geochemistry of redox sensitive trace metals in sediments. Geochim Cosmochim Acta 63:1735–1750. 76. Reijonen I, Metzler M, Hartikainen H (2016) Impact of soil pH and organic matter on the chemical bioavailability of vanadium species: The underlying basis for risk assessment. Environ Pollut 210:371–379. 77. Trefry JH, Metz S (1989) The role of hydrothermal precipitates in the geochemical cycling of vanadium. Nature 342:531–533. 78. Rubin K (1997) Degassing of metals and metalloids from erupting seamount and mid-ocean ridge volcanoes: Observations and predictions. Geochim Cosmochim Acta 61:3525–3542. 79. Sansone FJ, et al. (2002) Geochemistry of atmospheric aerosols generated from lava‐ seawater interactions. Geophys Res Lett 29:doi:10.1029/2001GL013882. 80. Whitfield M, Turner DR (1979) Water-rock partition coefficients and the composition of seawater and river water. Nature 278:132–137. 81. Duce RA, Hoffman GL (1976) Atmospheric vanadium transport to the ocean. Atmos Environ 10:989–996. 82. Andreae MO, Rosenfeld D (2008) Aerosol-cloud precipitation interactions. Part 1. The nature and source of cloud-active aerosols. Earth Sci Rev 89:13–41. 83. Mather TA, Pyle DM, Oppenheimer C (2003) Tropospheric volcanic aerosol. Volcanism and the Earth’s Atmosphere, Geophysical Monograph (Am Geophys Union, Washington, DC), Vol 119, pp 189–212. 84. Hinkley TK, Lamothe PJ, Wilson SA, Finnegan DL, Gerlach TM (1999) Metal emissions from Kilauea, and a suggested revision of the estimated worldwide metal output by quiescent degassing of volcanoes. Earth Planet Sci Lett 170:315–325. 85. Mather TA, et al. (2012) Halogens and trace metal emissions from the ongoing 2008 summit eruption of Kilauea volcano, Hawai’i. Geochim Cosmochim Acta 83: 292–323. 86. Mather TA (2015) Volcanoes and the environment: Lessons for understanding Earth’s past and future from studies of present-day volcanic emissions. J Volcanol Geotherm Res 304:160–179. 87. Bagnato E, et al. (2013) Scavenging of sulphur, halogens and trace metals by volcanic ash: The 2010 Eyjafjallajökull eruption. Geochim Cosmochim Acta 103:138–160. 88. Rauch JN, Pacyna JM (2009) Earth’s global Ag, Al, Cr, Cu, Fe, Ni, Pb and Zn cycles. Global Biogeochem Cycles 23:GB2001. 89. Randerson JT, Chen Y, van der Werf GR, Rogers BM, Morton DC (2012) Global burned area and biomass burning from small fires. J Geophys Res Biogeosci 117: G04012. 90. US Environmental Protection Agency (2017) Air pollutant emissions trends data. Available at https://www.epa.gov/air-emissions-inventories/air-pollutant-emissionstrends-data. Accessed July 4, 2017.
125. Conko KM, Rice KC, Kennedy MM (2004) Atmospheric wet deposition of trace elements to a suburban environment, Reston, Virginia. Atmos Environ 38:4025–4033. 126. Shimamura T, Wada T, Iwashita M, Takaku Y, Ohashi H (2007) Scavenging properties of major and trace species in rainfall collected in urban and suburban Tokyo. Atmos Environ 40:4220–4227. 127. Poissant L, Schmit JP, Beron P (1994) Trace inorganic elements in rainfall in the Montreal Island. Atmos Environ 28:339–346. 128. Moreda-Pineiro J, et al. (2014) Influence of marine, terrestrial and anthropogenic sources on ionic and metallic composition of rainwater at a suburban site (northwest coast of Spain). Atmos Environ 88:30–38. 129. Zhou J, et al. (2012) Origin and distribution of trace elements in high-elevation precipitation in southern China. Environ Sci Pollut Res Int 19:3389–3399.
130. Dasch JM, Wolff GT (1989) Trace inorganic species in precipitation and their potential use in source apportionment studies. Water Air Soil Pollut 43:401–412. 131. Heaton RW, Rahn KA, Lowenthal DH (1990) Determination of trace elements, including regional tracers, in Rhode Island precipitation. Atmos Environ 24:147–153. 132. Kaya G, Tuncel G (1997) Trace element and major ion composition of wet and dry deposition in Ankara, Turkey. Atmos Environ 31:3985–3998. 133. Al-Momani L (2003) Trace elements in atmospheric precipitation at northern Jordan measured by ICP-MS: Acidity and possible sources. Atmos Environ 37:4507–4515. 134. Hu GP, Balasubramanian R (2003) Wet deposition of trace metals in Singapore. Water Air Soil Pollut 144:285–300. 135. Baez A, Belmont R, Garcia R, Padilla H, Torres MC (2007) Chemical composition of rainwater collected at a southwest site of Mexico City, Mexico. Atmos Res 86:61–75.