Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs

Aeolian Research xxx (2013) xxx–xxx

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Review

Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs R. Scott Van Pelt ⇑ USDA-Agricultural Research Service, Wind Erosion and Water Conservation Research Unit, 302 W. I-20, Big Spring, TX 79720, USA

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Article history: Available online xxxx Keywords: Soil redistribution Wind erosion Radioisotopes Global fallout 137 Cs Spatial variability

a b s t r a c t Wind erosion is increasingly scrutinized as a causative factor in soil degradation and fugitive dust emissions. Although models have been developed to predict wind erosion and dust emissions, they are not accurate in all locations. The temporal and spatial variability of aeolian processes makes local estimates of long-term average erosion costly and time consuming. Atmospheric testing of nuclear weapons during the 1950s and 1960s resulted in anthropogenic radioisotopes that had not previously existed being injected into stratospheric global circulation and subsequently deposited on the Earth’s surface. Many of these radioisotopes are strongly adsorbed to soil particles and their movement on the landscape is a powerful method for investigating soil redistribution by wind, water, and tillage. 137Cs is the most commonly used anthropogenic radioisotope used to assess soil redistribution rates. Models have been developed to equate differences of radioisotope inventories with rates of soil redistribution and these models have been employed globally to assess soil redistribution on agricultural and natural landscapes. The radioisotope method for assessing soil redistribution rates has many advantages, but also a few limitations. One of the major limitations occurs when local sources of radioisotope contamination, particularly 137 Cs, mask the pulse from global fallout, making temporal estimates of redistribution difficult or impossible. In this paper, I explore the importance, history, and applications of the radioisotopic technique using 137Cs, particularly as it applies to soil redistribution by wind. Published by Elsevier B.V.

Contents 1. 2.

3.

4.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Nuclear age and atmospheric fallout . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1. Atmospheric testing of nuclear weapons. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2. Temporal patterns of global fallout . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3. Health and environmental concerns of radioactive fallout . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4. Environmental fate of anthropogenic radioisotopes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Radioisotopes and soil erosion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Early discoveries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. Model development . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.1. Cultivated soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.2. Undisturbed soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.3. Other models . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sampling and measuring soil-bound radionuclides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abbreviations: Am, americium; Ba, barium; Be, beryllium; Ca, calcium; Ce, cerium; Cs, cesium; I, iodine; IAEA, International Atomic Energy Agency; ICP/MS, inductively coupled plasma/mass spectrometry; K, potassium; LTER, long term ecological research; Np, neptunium; Nd, neodymium; NTS, Nevada test site; PR China, People’s Republic of China; Pb, lead; Pr, praesodymium; Pu, plutonium; Rh, rhodium; Ru, rubidium; SNAP, systems for nuclear auxiliary power; Sr, strontium; TNT, tri-nitro toluene; U, uranium; USA, United States of America; USDA, United States Department of Agriculture; USLE, Universal Soil Loss Equation; USSR, Union of Soviet Socialist Republics; WEPS, wind Erosion Prediction System; Zr, zirconium. ⇑ Tel.: +1 432/263 0293; fax: +1 432/263 3154. E-mail address: [email protected] 1875-9637/$ - see front matter Published by Elsevier B.V. http://dx.doi.org/10.1016/j.aeolia.2012.11.004

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

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5. 6. 7.

8.

9.

Advantages and limitations of the radioisotope technique . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Selection of reference areas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Wind erosion specific studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.1. Early discoveries and plot-scale research in Canada . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2. Other North American investigations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3. Southern Hemisphere investigations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.4. A holistic approach in the Sahel of Africa . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5. Asian investigations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Non-stratospheric fallout sources of anthropogenic radioisotopes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.1. Low yield and failed detonations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2. Accidental releases from peaceful nuclear activity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions and future prospects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1. Introduction In our modern anthropocentric world, wind erosion is viewed primarily as a soil degrading process. Wind erosion is often accelerated by human activity such as livestock grazing (Neff et al., 2005; Muminov et al., 2010), soil tillage (Sharratt et al., 2010), construction (Keating, 2003), and recreation (Goossens and Buck, 2009). Recently soil conservation position papers have listed soil erosion as one of the major threats to present and future sustainability of agricultural production (Delgado et al., 2011). Wind erosion removes the finer, more chemically active and nutrient rich portion of the soil (Zobeck and Fryrear, 1986; Van Pelt and Zobeck, 2007) and may adversely affect soil water dynamics (Lyles and Tatarko, 1986). Wind erosion also damages crops through sandblast injury (Baker, 2007) and dust deposition (Farmer, 1983). However, in some areas of the world, dust deposition results in plant nutrient inputs that are critical for ecological and agronomic sustainability (Poortinga et al., 2011). Finally, wind erosion results in fugitive dust emissions that negatively impact air quality (Sharratt and Lauer, 2006) but also helps to neutralize anthropogenic acids in clouds, resulting in less acidic rain downwind of industrial regions (Loye-Pilot et al., 1986). Soil may be eroded or redistributed by several forces other than wind and these forces rarely act alone. Detachment of soil particles by raindrop impact and entrainment of those particles into overland flow and runoff constitute water erosion, but often the coarser sand-sized particles are left behind for wind to entrain later (Zobeck, 1989). In a similar way, downslope and mechanical piling and movement of soil by tillage implements results in what is termed tillage erosion, but the bare fallow soil may be entrained by wind and moved further downslope or may even be moved back upslope by the wind. Wind erosion as discussed here is the entrainment and movement of individual soil particles or small aggregates by wind, air moving in a high velocity stream more or less parallel to the soil surface (Bagnold, 1941). Since the wind may blow from any direction, wind erosion does not move with gravity in a predictable direction as with water and tillage erosion and sediments may actually return to their point of origin on the landscape. In most cases however, coarse to medium sediments collect in nearby aerodynamically rough areas downwind of the source field (Hagen et al., 2007) and the fine fugitive dust that is the most visible evidence of wind erosion is lost from the source soil and transported long distances before returning to earth (Pye, 1987). Research over the last several decades has resulted in an understanding of the factors and processes controlling wind erosion. This body of knowledge has resulted in the creation of predictive models from the empirically-based Wind Erosion Equation (Woodruff and Siddoway, 1965) to the more process-based Wind Erosion Prediction System (WEPS) (Hagen, 2004). These models are used by

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land managers and government agencies such as the USDA–Natural Resources Conservation Service in the USA to evaluate potential land management effects on wind erosion. WEPS has been shown to match measured wind erosion reasonably well at several locations in the USA (Hagen, 2004) and very well in Germany (Funk et al., 2004). In spite of model utility in predicting wind erosion losses, they are sensitive to soil surface conditions such as residue cover, soil moisture, and crust cover (Feng and Sharratt, 2005) and their accuracy could benefit from local calibration (Van Pelt and Zobeck, 2004). Local calibration of models requires a locally derived measurement of wind erosion to compare with model output. Unfortunately, direct measurements of wind erosion are costly in terms of time due to its large temporal and spatial variability (Chepil and Woodruff, 1963; Chappell, 1999). Wind speeds and directions fluctuate on scales of seconds for individual events and the number of erosive events varies between months and years. Stroosnijder (2005) points out that, unlike measurements of water erosion, sediment cannot be collected at a single point because the wind blows from many directions, windblown particles occur at many heights above the ground, and the area of particle origin cannot be delineated. This necessitates the collection of a large number of spatially distributed samples on a frequent basis (Chappell et al., 2003). Clearly, a simple method of measuring long-term average wind erosion rates would facilitate local model calibration. The use of anthropogenic radioisotopes from atmospheric testing of nuclear weapons as tracers in soil has been shown to be an efficient method for locally calibrating predictive models of soil redistribution by water (He and Walling, 2003). This method assumes that fallout was uniformly deposited on the landscape within a local area, was adsorbed strongly and irreversibly to soil particles, and thus can only move about the landscape on the soil particles. In this article, I will introduce the origins of anthropogenic radioisotopes, review the discovery of their utility as erosion tracers, track the development of the method, and review the extent of its use globally. I will also discuss the advantages and limitations of the radioisotopic method, caveats for its proper use, and prospects for future use.

2. Nuclear age and atmospheric fallout 2.1. Atmospheric testing of nuclear weapons On July 16, 1945, at 5:30 in the morning Mountain Standard Time, the peace of the early dawn was shattered by a blast in the New Mexico, USA desert that was equivalent to 19,000 tons of TNT. This blast was quickly followed by two detonations over Japan in August and thus the world entered the nuclear age. The force in the bombs was created by the chain reaction of atomic nuclei fission and, in addition to unused nuclear fuel, many fission daughter

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

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Fig. 1. Relative rates of

137

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Cs fallout by year for the Northern and Southern Hemispheres (after Walling and Quine, 1990).

products were created which are distributed with the prevailing wind until they settle to the earth (Hala and Navratil, 2003). Most of the fission daughter products were short lived, but a few were longer lived and settled back to earth near and downwind of the blast sites, especially for low altitude explosions (Rabinowitch, 1956; Ballantyne, 1961; Beck and Krey, 1983; Simon et al., 2004; Saito-Kokubu et al., 2007). The anthropogenic radioisotopes 137Cs and 239+240Pu have been found in layers of arctic ice deposited in 1945 (Kudo et al., 1998). By the 1950s, other nations, including the USSR, Great Britain, and France, had developed nuclear weapons. Nuclear weapons testing and atmospheric fallout of radioactive fuel and daughter products increased as nations raced to develop and perfect nuclear weapons (Robbins, 1985). The US tested weapons in Nevada and in the northwestern tropical Pacific basin (Norris and Cochran, 1994) and the USSR tested weapons in northeastern Kazakhstan and over the Arctic Island of Novaya Zemlya (Ministry of Russian Federation for Atomic Energy, 1996). Except for a few detonations in the Southern Hemisphere by France and Great Britain, almost all atmospheric testing was done between 10° and 80° North latitude. As the yields of the devices became greater, the nations exploded most of the devices at altitude to reduce local fallout contamination, resulting in stratospheric contaminants that circled the globe several times allowing for the short-lived very radioactive species’ decay before fallout occurred (Simon et al., 2004). Although Northern Hemisphere fallout of anthropogenic radioisotopes is primarily from high yield polar region air burst testing and is very uniform in its composition, Southern Hemisphere fallout is largely tropospheric and from lower yielding devices and therefore has more variable composition (Kelley et al., 1999). 2.2. Temporal patterns of global fallout The Nuclear Test Ban Treaty of 1963 essentially halted atmospheric testing by the US and USSR, but in the months leading up to December 31, 1962, both nations were detonating multiple bombs each month, resulting in a peak loading of atmospheric fallout during the years of 1963 and 1964. By the end of atmospheric testing by the USA and USSR, 541 detonations had introduced approximately 6 Mg of 239Pu nuclear fuel into the environment (Warneke et al., 2002). The amount of fallout deposited on the Earth’s surface at a given location was a function of latitude and rainfall (Anderson, 1958; Davis, 1963) and, in the absence of new inputs, fallout decreased exponentially until the 1970s when atmospheric fallout had essentially ended (Robbins, 1985; Walling and He, 1999). Relative Northern and Southern Hemisphere fallout deposition rates for the years 1954–1985 are presented in Fig. 1. 2.3. Health and environmental concerns of radioactive fallout The Nuclear Test Ban Treaty was enacted because the nuclear nations of the world recognized the adverse environmental and health effects of radioactive fallout. As early as 1956, elevated gamma activ-

ities of people and foodstuffs had been reported and a detailed geographic study of people and dried milk samples in 1957 revealed that higher activities of 137Cs and 40K could be found in areas downwind of nuclear test sites and accidental releases (Anderson, 1958). The fission daughter products present in fallout included radioisotopes of U, Pu, Cs, I, and Sr that had previously not existed or had only existed in very small amounts in very localized conditions. Due to health concerns regarding the presence of these new isotopes, researchers began investigating their environmental fate. 2.4. Environmental fate of anthropogenic radioisotopes Immediately after the fission event, very short lived radioisotopes such as 131I and 140Ba produce the most radioactivity. After a few months, 141Ce, 95Zr, 95Nd, and 89Sr are the principal species undergoing fission and releasing radiation. On the timescale of 2–3 years, 144Ce, 144Pr, 106Ru, 106Rh, and 147Pr are the most radioactive species. After a few years, only non-critical mass fissile materials including 238U and 239+240Pu and relatively stable fission daughter species such as 90Sr, 129I, and 137Cs remained (Hala and Navratil, 2003). In many of the early fallout distribution studies, 90 Sr was the radioisotope of interest because of its ability to substitute for Ca in biological systems and its potential incorporation into bone tissue (Ballantyne, 1961). Unspent fissile material and fission daughter products vary in their environmental transport and fate. Gases such as 129I move freely in the environment even when in anion form (Edwards, 1962; Oktay et al., 2000). Unspent fissile species such as isotopes of U and Pu are bound to soil particles, but U is only about as rare as Mo or As in the environment and is often found in soil chemical analyses (Chemical Rubber Company, 1985; Litaor and Ibrahim, 1996; Beasley et al., 1998). Long-lived fission daughters 90Sr and 137 Cs are also bound tightly to soil particles (Schulz et al., 1960; Tamura, 1964; He and Walling, 1996; Quang et al., 2004; Bossew et al., 2007; Bihari and Dezso, 2008) and soil organic matter. So strong is the correlation between 137Cs and soil organic matter that it has been used to track soil organic matter on the landscape (McCarty and Ritchie, 2002; Ritchie et al., 2007). Parsons and Foster (2011) question the affinity of 137Cs for soil particles and organic matter, however the predominance of evidence would indicate otherwise. Other fission daughter products also attach to the soil particles but may be removed from the soil solution and create a directional gradient resulting in non-decay and non-erosional decreases of inventory with time. 90Sr is a biological analog for Ca, is taken up by plants through the roots, and enters the food chain (Mouat, 1960; Vose and Koontz, 1960; Haghiri, 1964). Conversely, 137 Cs is not readily taken up by plant roots and only enters the food chain when rain splash or wind erosion results in soil particles becoming attached to the ingested plant (Beresford and Howard, 1991; Papastefanou et al., 1999; Muminov et al., 2010). Copplestone et al. (2001) credit ‘bursting bubbles’ of turbid coastal water returning eroded anthropogenic radioisotopes to the shoreline as an additional pathway of vegetation contamination.

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

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Parsons and Foster (2011) dispute the claim that 137Cs is not taken up by plants and suggest that as much as 1/8 of the original inventory may have been intercepted by plants before reaching the soil. They also cite studies in which such factors as mycorrhizal infection and soilborne chelates or ammonium and potassium ions may influence the plant absorption of 137Cs. In general however, studies in which field harvested plants were compared with respect to washed vs. unwashed have established that in soils from semi-arid regions low in organic matter, the vast majority of 137 Cs is attached to the outside of the plant (Muminov et al., 2010). Infiltration of fallout bearing rainwater and incorporation of dry deposited atmospheric fallout results in incorporation of the radioisotopes into the soil. Although infiltration is often assumed to be a spatially uniform process, it is in fact not uniform at all. Surface cracks and macropores resulting from soil faunal activity represent sites in which very rapid influx of rainwater may occur to various depths (Das Gupta et al., 2006). This spatial variability has been documented in undisturbed soils (Wilson and Luxmoore, 1988; Timlin et al., 1994) often chosen to quantify the total radioisotope input from fallout. Other patterns of spatial variability have been documented in a cultivated field soil (Van Pelt and Wierenga, 2001). Tillage may add a temporal variability of infiltration rates during and between seasons (Logsdon and Jaynes, 1996). In addition to the spatial variability of infiltration rates, microtopography also resulted in differences of radioisotope input as rainfall in excess of infiltration rates ran off the relatively high ground and ponded in the microdepressions (Kachanoski et al., 1985; Kachanoski and de Jong, 1988). Wind-driven drifting of snow may also have added another component to the spatial variability of radioisotope inputs, but due to the time of year, de Jong et al. (1982) estimate that this would only cause a 7% difference. Parsons and Foster (2011) point to all these sources of variability as problems with the radioisotopic method for estimating erosion rates and, further, cite evidence of spatially variable deposition measured during the period following the nuclear power plant fire in Chernobyl, Ukraine and to spatially variable rainfall measurements at the Walnut Gulch experimental wa-

tershed in Tombstone, Arizona, USA during the period of global fallout deposition from 1956 to 1979. We may thus surmise that 137 Cs inventory is highly variable and only represents the incorporation and subsequent redistribution history of that point on the landscape. Due to the spatial variability inherent in precipitation infiltration, sampling should include as many points on the landscape as possible. Where time and budget allows, the samples may be analyzed for 137Cs inventory individually, resulting in a data set from which geostatistical parameters may be inferred. The use of a nested sampling strategy where a tighter grid of sampling sites is situated in each distinct soil type or geomorphologic feature and a looser grid covers the landscape has been shown to be efficient and effective for improving accuracy (Chappell, 1999; Chappell and Warren, 2003). When time and budget are limiting, samples from many points on the landscape may also be composited by depth, resulting in a single sample profile for analysis which represents the areal mean 137Cs inventory. 137 Cs was also deposited on the landscape on dry particles of dust where protection from re-entrainment by wind allowed for eventual incorporation into the soil. Dry deposition may be less spatially variable than wet deposition except where abrupt changes in vegetation canopies occur. This phenomenon has been demonstrated at forest edges where tree canopy filtering of air masses and subsequent precipitation throughfall washing of foliage results in higher soil radioisotope inventories than adjacent areas covered with grass or heath (Bunzl and Kracke, 1988; Branford et al., 2004). In a study of attic dust in New Jersey, USA, Ilacqua et al. (2003) found that the activity of 137Cs in attic dust was greater for homes that existed in the period of atmospheric fallout than in more recent structure, again indicating that dry deposition of fine particulate was a significant pathway of radioisotope deposition on the landscape. Since the long-lived fission daughters such as 137Cs are strongly adsorbed to the soil particles, they are found primarily near the surface of the soil profile and tend to form a profile of activity that decreases logarithmically with depth (Ritchie and McHenry, 1990; He and Walling, 1997; Ritchie, 2000; Sigurgeisson et al., 2005;

Fig. 2. Idealized profiles for an area receiving 1575 Bq m2 of atmospheric fallout. Subfigure a represents an undisturbed profile without erosion or deposition; subfigure b represents an undisturbed profile with an eroded surface; subfigure c represents a cultivated soil with some minor erosion; subfigure d represents an undisturbed soil with 10 cm of post-fallout deposition.

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

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Al-Masri, 2006; Whicker and Ibrahim, 2006; Osaki et al., 2007). In undisturbed soils with no post-deposition erosion or deposition history representing an ideal reference site, the greatest 137Cs activity is found just below the surface and rapidly decreases with depth (Fig. 2a). In an untilled soil that has been eroded, the peak concentration may be removed and the total activity reduced compared with the reference site (Fig. 2b). In a tilled soil, the peak is less distinct due to the incorporation of unlabeled soil from below (Fig. 2c) and the status of erosion or deposition is determined primarily from the total activity in the profile or, when expressed as a function of surface area, the 137Cs inventory for that soil. Undisturbed soils on which deposition has occurred often have a buried peak (Fig. 2d). The exact nature of the undisturbed profile and ultimate fate of the incorporated 137Cs depend on the depositional environment and vegetation community present. Thick surface litter layers under forested locations may intercept the fallout radioisotopes and, if burned, result in the rapid release of organically-bound fission daughters in surface runoff and wind dispersion of ash (Johansen et al., 2003). In lacustrine sediments and soil profiles that receive sediment deposition containing adsorbed 137Cs on the particle surfaces, the peak concentrations from the initial infiltration during the years of peak fallout may become buried and, in the northern hemisphere, offer a horizon marker of the surface in 1963–1964 (Ritchie and McHenry, 1990; Oktay et al., 2000; Copplestone et al., 2001; Vanden Bygaart and Protz, 2001; Michel et al., 2002; Yan et al., 2002; Li et al., 2003). In coarse soils, the radioisotopes tend to percolate to a greater depth than in soils with clay and organic matter (Vanden Bygaart and Protz, 2001; Quang et al., 2004). With time, illuviation of clays and large organic molecules and soil movement by soil fauna will broaden the peaks found in the profile (Ritchie and McHenry, 1990; Muller-Lemans and van Dorp, 1996; Sigurgeisson et al., 2005; Whicker and Ibrahim, 2006).

3. Radioisotopes and soil erosion 3.1. Early discoveries The relationship between anthropogenic radioisotopes from atmospheric fallout and soil erosion was first reported in 1960 when a USDA researcher was monitoring runoff from standard erosion plots for 90Sr. Menzel (1960) discovered that the transport of fallout 90Sr was always greatest on plots where the most soil erosion occurred. He also found that the plots with the least runoff and erosion had the greatest activity of 90Sr. A few years later, the work of Frere and Roberts (1963) confirmed the relationship of soil redistribution with 90Sr content of the runoff from erosion plots in Ohio. Subsequent investigations were able to confirm the relationship between radioisotope transport and soil particle redistribution by artificially adding radioisotopes 85Sr and 131I (Graham, 1963) and 137Cs (Rogowski and Tamura, 1965, 1970a,b) to small catchments and monitoring the runoff. Rogowski and Tamura (1970b) concluded that erosion was a major factor on the redistribution of 137Cs on the landscape and suggested an exponential model equating 137Cs activity with soil loss. In the late 1960s, Ritchie and McHenry began a series of experiments to determine if fallout levels of 137Cs could be used to determine rates of soil movement across natural and agricultural landscapes (Ritchie, 1998). They reported a logarithmic relationship between computed soil loss using the Universal Soil Loss Equation (USLE) (Wischmeier and Smith, 1965) for three north Mississippi watersheds and concluded that although it was unlikely that one equation would fit radionuclide loss and soil erosion at all locations, there was probably a family of logarithmic equa-

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tions that would work in most cases (Ritchie et al., 1974). Ritchie and McHenry (1975) combined their data with the previous findings of Menzel (1960), Graham (1963), Frere and Roberts (1963), and Rogowski and Tamura (1970b) and again found a significant exponential relationship between soil loss and radioisotope loss. During the 1970s, most of the radioisotope and soil erosion research was in the United States but by the early 1980s, there was ongoing research in Australia, Great Britain, and Canada as well (Ritchie, 1998). Much of this work was reviewed by Ritchie and McHenry (1990) and later documented in an exhaustive bibliography of 137Cs and soil-related research (Ritchie and Ritchie, 1995, 2007). 3.2. Model development As the body of evidence of the relationship between 137Cs activity and soil loss accumulated, models evolved that allow computation of soil redistribution rates on decadal timescales based on measured 137Cs soil inventory compared with the 137Cs inventory expected if neither erosion nor deposition had occurred at the site. The amount of 137Cs present on a mass basis is termed 137Cs activity and is expressed in units of Bq kg1. The amount of fallout 137Cs received on a soil areal basis is termed 137Cs inventory and is expressed in units of Bq m2. Most models relating 137Cs redistribution with soil redistribution work with 137Cs inventories as input parameters. 3.2.1. Cultivated soils 3.2.1.1. The proportional model. The earliest and simplest form of the model was the linear proportion model (de Jong et al., 1983; Kachanoski and de Jong, 1984; Kachanoski, 1987; Martz and de Jong, 1987). This model is expressed as:

R ¼ 10

Bd

Aref A Aref

ðt  1963Þ

ð1Þ

where R, mean annual soil loss (t ha1 yr1); B, bulk density of the soil (kg m3); d, depth of tillage or cultivation layer (m); Aref, the 137 Cs inventory of an uneroded reference site (Bq m2); A, the 137 Cs inventory of the eroded site (Bq m2); t, the year of sample collection. This model is considered appropriate for cultivated soils in which the entire 137Cs inventory in the profile has been thoroughly mixed throughout the depths of cultivation. This simple model assumes that all the 137Cs arrived in 1963, the year in which most of the fallout was deposited on the surface in the northern hemisphere (the date of maximum fallout lagged by a year or so in the southern hemisphere and is more difficultly determined), the activity of 137Cs at the surface available for erosion remains relatively constant through time, and that the loss or gain of 137Cs in the profile is linearly and directly proportional to the loss of soil in the profile. For areas of sediment deposition, this model is expressed as:

R0 ¼ 10

Bd

AAref Aref

ðt  1963Þ

ð2Þ

where R0 , mean annual sediment deposition (t ha1 yr1). One of the limitations of the linear proportional model is that it assumes equal movement of 137Cs on a soil mass basis without consideration of the surface area of the particles on which the soil is adsorbed. He and Walling (1996) proposed a particle size correction factor, P, placed as a multiplier of (t  1963) in the denominator. The particle size correction factor takes into account the specific surface area of the mobilized sediment relative to that of the source soil. The particle size correction factor is expressed as:

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

6

P¼ð

R.S. Van Pelt / Aeolian Research xxx (2013) xxx–xxx

Sms v Þ Ssl

ð3Þ

where P, particle size correction parameter (unitless); Sms, specific surface area of mobilized sediment (m2 g1); Ssl, specific surface area of the source soil (m2 g1); v, a constant with a value of 0.65. From Eq. (3) it is apparent that if the displaced particles eroded from a soil or deposited on a surface were finer than the bulk soil of origin, P would have a value greater than 1. Conversely, if the displaced particles were coarser than the bulk soil of origin, P would have a value less than unity. Unfortunately, the difficulty of making an actual determination of P over time has resulted in some researchers ignoring this important parameter and using unity for its value in modeling efforts.

If we assume that the total 137Cs fallout occurred and that the depth distribution of the profile is time invariant, then the mean annual erosion rate can be expressed as the simple profile distribution model:

R¼

Aref  Au 10 lnð1  Þa ðt  1963ÞP Aref

where Au, measured total 137Cs inventory at the sampling point (Bq m2). For a location that is accreting, the depositional rate R0 can be estimated from the following:

R0 ¼ R t t0

3.2.1.2. Mass balance models. In cultivated soils, tillage operations tend to mix 137Cs deficient soil from below the zone of incorporation in the profile to replace surface soil lost to erosion. Zhang et al. (1990) proposed a mass balance model that would account for this mixing. In addition, although atmospheric fallout peaked in 1963 in the northern hemisphere, it occurred from 1954 until the early 1980s. This results in the need for a time variant rate of 137Cs input in the model. For locations without a detailed history of fallout activities, latitude and rainfall may be used as a proxy to estimate the total 137Cs input (Walling et al., 2006). The mass balance model was further refined by incorporating the time variant rate of 137Cs addition, a parameter to account for the loss of 137Cs prior to incorporation by tillage (He and Walling, 1997), and a particle size correction factor. This improved mass balance model (Walling and He, 1999) is expressed as: PR

AðtÞ ¼ Aðt 0 ÞeðdþkÞðtt0 Þ þ

Z

t

R

PR

0

ð1  Pcð1  eðHÞ ÞÞIðt 0 ÞeðdþkÞðtt Þ dt

0

ð4Þ

t0

where A, 137Cs inventory at the time indicated (Bq m2); t, year of sampling; t0, year cultivation began; t0 , year that the 137Cs deposition began (generally1954 in the northern hemisphere); d, cumulative mass depth representing the average cultivation depth (kg m2); k, decay constant for 137Cs (yr1); I(t0 ), annual 137Cs deposition flux (Bq m2 yr1); P, particle size correction factor (unitless); c, proportion of the annual 137Cs flux susceptible to loss by erosion before incorporation into the soil profile; H, is the relaxation mass depth of the initial distribution of fallout 137Cs in the soil profile (kg m2). A(t0) may be estimated by:

Aðt 0 Þ ¼

Z

t0

0

Iðt0 Þekðt t0 Þ dt

0

ð5Þ

1954

Eq. (4) is solved numerically to yield an estimate of R. Tillage also moves surface soil and affects the 137Cs and erosion relationship if not taken into account (Quine, 1999). The improved mass balance model has been further modified to include surface soil redistribution by tillage (Walling and He, 1999). This refined model has been shown to reduce estimates of soil loss from 84 to 50 Mg ha1 yr1 for agricultural fields in Morocco (Nouira et al., 2003). 3.2.2. Undisturbed soils 3.2.2.1. Profile distribution model. As mentioned earlier, 137Cs is strongly adsorbed to soil mineral and organic matter and the profile distribution of its activity decreases with depth. This exponential decline with depth has been expressed as (Zhang et al., 1990; Walling and Quine, 1990):

Ax ¼ Aref ð1  ex=a Þ

ð6Þ 137

where Ax, amount of Cs above a given depth (x cm) in the profile (Bq m2); a, coefficient describing profile shape.

ð7Þ

Aex C d ðt

0

0 0 Þekðtt Þ dt

¼

R P0 S

RdS

R

Au  Aref A ð1  eR=a ÞdS S ref

ð8Þ

where Aex, the excess 137Cs present in the profile due to deposition = Au  Aref (Bq m2); Cd(t0 ), the 137Cs concentration of deposited sediment (Bq kg1); P0 , a specialized particle size correction factor similar to that in Eq. (3) relating the surface area of the deposited sediment with that of the mobilized sediment; R, the erosion rate of the contributing area (kg m2 yr1); S, the size of the sediment contributing area (m2). 3.2.2.2. Diffusion and migration model. For reasons of time variant 137 Cs fallout and redistribution of 137Cs within the profile (He and Walling, 1997), the simple profile distribution model above is not entirely accurate (Walling and He, 1999). Walling and He (1999) have proposed a much more complex model for undisturbed soils, the diffusion and migration model, that takes into account the complex time-dependent 137Cs deposition and redistribution processes. Both the simple profile distribution and the diffusion and migration models have been widely used in studies of soil redistribution on uncultivated soils (Ritchie et al., 2003, 2005). All of the models discussed above have been compiled and made readily usable through a computer spreadsheet interface (Walling et al., 2006). 3.2.3. Other models In addition to the models mentioned above, other researchers have proposed models relating soil redistribution rates to radionuclide activity. Brown et al. (1981) compared a volumetric approach with a gravimetric approach based on differences of 137Cs measured on agricultural landscapes on Oregon and found, that in spite of how crude both approaches were, they yielded very similar results. A process based model has been proposed that iterates on a yearly time step to account for inter-annual variations in fallout and surface management. This model takes into account rill development and transport capacity, localized deposition, tillage, and root harvest erosion (Van Oost et al., 2003). Chappell and Warren (2003) investigated the spatial distribution of 137Cs on the landscape and implications for the use of mass balance models to determine soil redistribution by wind. They modified a mass balance model to operate on a daily time step and included erodibility and erosivity as input parameters so temporal changes in soil redistribution and land management could be used to predict spatial changes in 137Cs inventory. Heuvelink and Webster (2001) have proposed the use of the Kalman filter to improve the spatial accuracy of erosion rate estimates between actual sampling points. A mass balance model for use with undisturbed soil profiles has also been proposed and validated with field measurements for slopes in a Mediterranean climate (Soto and Navas, 2004). More recently, Zhang et al. (2008) have proposed a simplified model for undisturbed profiles that fits a dispersion coefficient and redistributes the 137Cs downward through the profile on an annual basis. Their model results in smaller estimates of soil

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

R.S. Van Pelt / Aeolian Research xxx (2013) xxx–xxx

redistribution than the commonly used profile distribution model presented above. 4. Sampling and measuring soil-bound radionuclides In order to quantify the redistribution of radionuclides, the soil profiles must be sampled at multiple locations on the landscape (Ritchie and McHenry, 1990; Walling, 1998). For deeply tilled soils, a single soil core taken to below the depth of tillage may be mixed, a subsample removed and weighed, the subsample dried and corrections made to the original weight of soil sampled at location. In the laboratory, a 100–1000 g dry soil sample is placed in a Marinelli beaker, and the activity measured in a suitable instrument, most commonly a gamma spectrometer with Ge detector coupled with a multi-channel analyzer. For the commonly measured 137Cs, activity is estimated using 662 keV peak (Murray et al., 1987; Ritchie and McHenry, 1990; Wallbrink and Murray, 1996; Schuller et al., 2007; Muminov et al., 2010). For undisturbed soils, the distribution of the nuclide with depth is essential information. For sampling with split-spoon coring devices that obtain undisturbed cores, the core may be cut at selected depths and the discrete depth samples from multiple cores composited for gamma spectrometry (de Jong et al., 1982). For a single point on the landscape, the scraper plate method may be used where the edges of a rectangular area are set within a surface mounted frame and soil is removed by layers using a depth-adjustable scraper that rides along the frame (Loughran et al., 1987; Krause et al., 2003; Funk et al., 2011). In some cases, pits have been dug and samples take from discrete depths along the pit wall (Ritchie et al., 2003). More recently, in situ measurements of 137 Cs activity have been made using a field portable gamma detector mounted on a tripod above the soil (He and Walling, 2000; Li et al., 2010; Funk et al., 2011). 5. Advantages and limitations of the radioisotope technique Using anthropogenic radioisotopes to estimate rates of soil redistribution offers many advantages (Walling and Quine, 1991; Walling, 1998; Zapata, 2003; Mabit et al., 2008).  Estimates are based on the date of sampling and provide retrospective estimate of multi-decadal (50 years) rates of soil redistribution.  Estimates can be obtained with a single site visit.  Estimates are integrated and less influenced by extreme events compared to multi-years sampling studies.  Estimates are from single points on the landscape allowing information on both rates and spatial patterns on soil redistribution.  Sampling does not require significant disturbance of the study area.  The results are compatible with physically based distributed modeling, geostatistics, and GIS techniques used to study soil redistribution and can be used for local calibration of predictive models.  Estimates represent integration of all soil redistributing forces including water, wind, and tillage.  The technique provides erosion and deposition information for a single land unit, allowing estimates of net loss or gain.  There are no scale-based constraints other that the number of samples to be analyzed.  The technique allows quantification of soil loss due to sheet erosion.  The technique allows ongoing monitoring of soil redistribution by sampling at the same sites on sequential date, usually separated by years.

7

No technique is perfect and there are recognized limitations as, well.  The complexity of the technique requires an interdisciplinary team and specialized equipment that may not be available in developing countries.  In some areas, especially the southern hemisphere or areas with very low rainfall, the small amounts of anthropogenic radioisotopes present in the soil requires longer counting times and quality assurance measures.  The technique is essentially limited to surface erosion and does not provide estimates of gully erosion or of erosion beyond the depth of radioisotope incorporation in the profile.  It is an indirect approach that depends on the link between measured soil redistribution and observed radioisotope redistribution.  There are uncertainties associated with the selection and use of the conversion models for converting measurements of radioisotope inventories to estimates of soil redistribution rates.  The technique does not allow short-term estimates of rates for the study of changes in soil redistribution rates due to shortterm changes in land management.  The technique could benefit from standardized protocols that would allow comparison of rates derived from different studies.  The technique is sensitive to local inputs of radioisotopes from industrial contamination or accidents.  The technique is sensitive to microtopographic influences that would have affected runoff and infiltration during the years of fallout deposition.  The technique is dependent on the inventory obtained from an undisturbed reference site that has undergone neither erosion nor deposition. One advantage that does not explicitly appear in the list above is the cost to obtain the estimates. Obviously, a single site visit will be much, much less expensive that a multi-year field campaign with multiple site visits each year. In addition, the multi-decadal period of rate estimation integrates the temporal variation in erosion rates that might not be adequately sampled in a field measurement campaign.

6. Selection of reference areas The last limitation mentioned above has been the source of considerable debate and inquiry. Considering the 50 years that have passed since fallout started accumulating in the soil, it is difficult to determine that a given site has not been disturbed, eroded, or sedimented during the entire time. Cemeteries, airports, parks, schoolyards, lawns, well vegetated livestock exclosures and other areas established prior to 1954 that have been administered and have a record of operations are good candidates for reference sites (Basher et al., 1995; Walling, 1998; Ritchie, 2000; Ritchie et al., 2003, 2005). The areas should be flat to limit water movement and associated erosive forces and well vegetated to protect the surface from erosion be either wind or water (Porto et al., 2003; Funk et al., 2011). There are many factors that control the spatial variability of radioisotope content of soils in any given region. Some investigators claim that variability of precipitation infiltration caused by circulation patterns, rain shadowing, runon/runoff processes, and snow drift at the time of radioisotope deposition may affect the spatial patterns (de Jong et al., 1982; Kiss et al., 1988; Wallbrink and Murray, 1996; Parsons and Foster, 2011). It has been recently suggested that in skeletal soils on steep slopes, the use of radioisotopes to estimate redistribution rates may be of limited utility due

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

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R.S. Van Pelt / Aeolian Research xxx (2013) xxx–xxx

to rapid runoff and limited incorporation of radioisotopes during the years of fallout deposition (Zhang et al., 2011). Other investigators have found that the variability of radioisotope inventories in sites meeting the requirements for a good reference site are random and not related to landscape position (Pennock et al., 1994; Sutherland, 1994; Owens and Walling, 1996). They further recommended sampling 9–16 sites in the reference area to calculate the mean reference inventory (Pennock et al., 1994; Sutherland, 1994) and to express the reference inventory in terms of a mean and confidence intervals (Owens and Walling, 1996). In dusty locations typical of arid and semi-arid regions, it is particularly difficult to find sites without dust deposition (Chappell, 1999). This is particularly problematic due to the resuspension of radioisotopes with the emissions of aeolian dust from wind eroding soil surfaces (Papastefanou et al., 2001; Igarashi et al., 2001, 2003; Hirose et al., 2004; Arimoto et al., 2005; Choi et al., 2006; Karlsson et al., 2008; Chamizo et al., 2010). When the long-term rate of dust deposition is known, Chappell (1999) recommends multiplying that depth by the number of years since peak fallout deposition and removing the inventory represented by that layer from the reference inventory value. Funk et al. (2011) recommend using flat, well vegetated surfaces near the top of topographic highs and areas downwind of those highs in order to minimize dust deposition in keeping with the flow and deposition dynamics findings of Goossens (1988, 1996, 1997). One method to reduce the dependency on the reference inventory is to resample an area after a discrete period of time (Kachanoski and de Jong, 1984). This approach has not often been used, but it has the added advantage of assessing erosion for a shorter period than the five decades since the deposition peaked. If this approach is utilized, the investigator should be careful to take a large number of samples in order to lessen the error due to spatial variability and avoid the exact locations of the previous sample collection on the date of the second sampling. 7. Wind erosion specific studies All of the literature written on the link between fallout radioisotope redistribution and soil redistribution before the early 1980s was based on water erosion research. Indeed, most of the radioisotope redistribution literature up to the present time is water erosion based in a pattern similar to the fluvial-dominated soil redistribution literature in general (Field et al., 2009). Even the oft cited review by Ritchie and McHenry (1990) only mentions wind once and that occurrence is in parentheses after water. It would take nearly 20 years from the initial discovery of anthropogenic radioisotopes moving on eroded soil particles until the first mention of wind as a redistribution agent would be published in radioisotope redistribution literature (de Jong et al., 1982). 7.1. Early discoveries and plot-scale research in Canada De Jong et al. (1982) reported reference inventories on native grassland sites of 2183 ± 259 to 2405 ± 333 Bq m2. They observed windblown material near the edge of one of their fields and measured an increase of 137Cs inventory relative to undisturbed sites and predictions based on regionally measured 137Cs fallout. They also found 137Cs at a greater depth than in the rest of the study sites, indicating burial of the 1963 surface. Subsequently, a study of eight adjacent small closed basins in Saskatchewan, five of which were in cultivation, one in native prairie, and two had been in seeded pasture since the late 1950s (de Jong et al., 1983) mentioned wind erosion and the Wind Erosion Equation (Woodruff and Siddoway, 1965). Their results focused on slope position and 137 Cs inventory, commensurate with their primary interest in

water erosion. The reported mean137Cs reference inventories ranging from 2183 ± 555 to 2553 ± 962 Bq m2 in the undisturbed prairie sites. They concluded that in most of their cultivated study fields, 137Cs inventory was more strongly and inversely correlated with site elevation, indicating a downhill redistribution pattern expected with water erosion and sedimentation. However, they also reported apparent burial of a 137Cs rich surface horizon in a depression of the native prairie site that they postulated could be attributed to deposition of wind-eroded soil and also noted that the snow in the depressions contained soil material. Their 137Cs inventory-based estimate of erosion at the tops and upper slopes of knolls in cultivated fields did not agree with predictions made using USLE and they attributed this discrepancy to loss by wind erosion. The first investigation that focused on using 137Cs to estimate rates of soil redistribution primarily by wind and to contain wind in the title was published in 1991 (Sutherland et al., 1991). In this study the authors investigated soil redistribution rates on five nearly level fields with slopes <2% in southern Saskatchewan that were judged to be relatively immune to water erosion due to their sub-critical slopes. They used a linear proportion model, assumed that 95% of the fallout 137Cs was incorporated into the soil, and reported a mean reference 137Cs activity of 2110 ± Bq m2. Net 30 year mean sediment outputs for these fields ranged from 0.8 to 38.2 Mg ha1 yr1 and critical erosion values of 1.0, 5.0, and 11.0 Mg ha1 yr1 were exceeded on 77%, 71%, and 57% of the individual sample locations. They reported that soils with textures in the sandy loam and loam range eroded more readily than did soils in the silt loam texture class, probably due to the greater amount of readily entrained sand for saltation and the relative paucity of stable aggregates. The fields with silt loam soils also had the greatest number of sampling sites displaying net soil deposition. Southern Saskatchewan continued to be a region of intense scientific investigation of using anthropogenic radioisotopes to estimate rates of soil redistribution by wind. Sutherland (1994) investigated the spatial variability of 137Cs on the landscape and found, based on the 36 sampling points, that the reference site inventory was 2040 ± 120 Bq m2. In an adjacent cultivated field, he found inventories ranging from 950 to 3500 Bq m2 and attributed the differences outside the confidence intervals of reference inventory to soil redistribution by forces of primarily wind and, in localized cases, of water. He further noted that the localized nature of water erosion within the field will be complicated by the influence of slope-independent wind erosion and minor soil redistribution by tillage. Pennock et al. (1995) investigated the influence of soil parent material on decadal erosion rates based on the 137Cs technique in southwestern Saskatchewan in four cultivated fields and 1 uncultivated field for each of 5 parent material groups. They reported variability of median 137Cs reference inventories ranging from 1698 to 2373 Bq m2 among soil parent material groups, but they could not separate the median values because of in-field variability. In the cultivated fields, they also found great variability, but concluded that the soils derived from glacial till were more erodible and the soils derived from fine sandy loam and silty glaciolacustrine and aeolian material were less erodible than the soils derived from coarse sand glacio-fluvial-lacustrine materials.

7.2. Other North American investigations The 137Cs technique was employed to study soil redistribution processes in a sagebrush steppe in Wyoming, USA (Coppinger et al., 1991). They found significantly more 137Cs under shrubs in windswept landscape positions and attributed wind as the motive force causing the post-fallout movement of soil. They further noted that analysis of the soil texture in downslope transects indicated

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

R.S. Van Pelt / Aeolian Research xxx (2013) xxx–xxx

fluvial transport as well, but that the process was too slow to be determined by redistribution of 137Cs. Ritchie et al. (2003) studied the patterns of soil redistribution in several plant communities in southern New Mexico using 137Cs as a soil tracer. In this detailed investigation, the authors reported soil redistribution in all the plant communities sampled, but noted that in desert grassland, this redistribution was limited to the interplant scale and that these communities retained a reasonably homogenous distribution of 137Cs. They used Black Grama (Bouteloua eriopoda Torr.) grassland in an exclosure ungrazed for 60 years as their reference site. The reference inventory was reported as 1112 Bq m2 and, in other plant communities, they found 137Cs inventories ranging from 0 to 5333 Bq m2. Tobosa grass (Pleuraphis mutica Buckley), a bunch grass, dominated sites were found to have a relative uniform distribution of 137Cs, but net deposition from aeolian and fluvial sources was estimated at 5 Mg ha1 yr1 based on the profile model. At shrub dominated sites, the rates of soil redistribution were found to be greater with 137Cs inventories sampled under 3 individual tarbush (Flourensia cernua DC.) shrubs of 0, 745, and 4667 Bq m2. They could not quantify the rate here due to the total lack of radioisotope at one of the sites and attributed the much greater inventory noted in the one case as due to the presence of a well developed lichen crust at that location. At coppice dune sites dominated by Mesquite (Prosopsis glandulosa Torr.), they always found greater inventories of 137Cs under the mesquite canopies within the dune representing net deposition rates between 0.7 and 10.4 Mg ha1 yr1 and lesser inventories in the interdune blowout areas representing erosion rates of 3.2 to 4.1 Mg ha1 yr1, indicating a net movement from the blowouts to the dunes. In general they found a gradient of 137Cs inventories in these sites increasing in the downwind direction indicating redistribution by wind on a large landscape scale. The 137Cs technique has been used estimate rates of dune growth in coastal (Vanden Bygaart and Protz, 2001) and continental (Hall et al., 2010) locations. 7.3. Southern Hemisphere investigations Almost all the atmospheric testing of Nuclear weapons occurred at mid and high latitudes in the Northern Hemisphere resulting in relatively low levels of anthropogenic radioisotope fallout in the Southern Hemisphere (Fig. 1). In spite of the expected low reference inventories of anthropogenic radioisotopes, the technique has been successfully employed at locations in Australia and New Zealand. In the Nullarbor Plain of southern Australia, Gillieson et al. (1994) found that 137Cs inventories were greater under tarbush (Atriplex spp.) canopies than in inter-canopy area, indicating aeolian redistribution of soil from the open areas to the shrub canopies in a pattern consistent with the findings of North American researchers. However, they further note that lithogenic radioisotopes 226Ra and 232Th showed a much stronger parallel trend and postulated that aeolian transport rates may have been greater in the past than in the post-fallout period. In subsequent contributions, they report a reference 137Cs inventory of 450 ± 70 Bq m2 and state that redistribution is almost totally by wind and that the shrubs on the ridges filter dust from the air (Gillieson et al., 1996a) and that although there is moderate redistribution, there is little net loss (Gillieson et al., 1996b). In Western Australia, Harper and Gilkes (1994) evaluated the 137 Cs technique for estimating soil redistribution in sandy soils. They reported a reference 137Cs inventory from uneroded sites of 421 ± 26 Bq m2, very similar to the reference inventory used by Gillieson et al. (1996a) but with less than half the spatial variability. In the sandy soils, they also found 137Cs below the zone of cultivation and concluded that some leaching was occurring due to low clay contents in the soils and the presence of kaolinite with

9

a low cation exchange capacity. Because of the variability they found among reference samples, they advocate obtaining very large numbers of samples and, due to the cost of processing the samples and the uncertainties involved, they question the utility of the method for estimating marginal rates of erosion. More recently, geostatistical techniques have been employed to reduce the uncertainty of the 137Cs reference inventory (Chappell et al., in press-a) and of a previous nationwide estimate of soil redistribution rates for the period of 1954 to 1990 (Chappell et al., in press-b). In New Zealand, the 137Cs technique was used to assess the patterns of soil redistribution in two South Canterbury downlands fields with long term histories of either cropping or occasionally renewed pasture (Basher et al., 1995). The authors used the linear proportion model and a mean reference inventory of 627 ± 31 Bq m2 to convert 137Cs inventories to estimates of soil redistribution. They found that although variability of 137Cs inventories in the fields indicated considerable redistribution within the fields, there was little evidence of net loss on the cropped field. They noted however a pattern of slight deposition in the pasture suggesting aeolian deposition of eroded sediment. The senior author subsequently studied soil redistribution on native vegetated terraces at a more upland setting of the South Island’s Mackenzie Basin. They used a protected area of red tussock (Hieracium pilosella) grassland for the reference site and obtained a mean 137Cs inventory of 388 ± 25 Bq m2. For unprotected areas on the terraces, 137Cs activities ranged from 118 to 845 Bq m2 with the lesser values found on bare areas and the greater values in grass tussocks. Overall, they concluded that soil loss rates were negatively correlated to vegetative cover and overall soil loss was 2.2 cm and non-tussock soil loss was 2.8 cm over the 30 year period. 7.4. A holistic approach in the Sahel of Africa One of the more holistic investigations of landscape scale soil redistribution by wind was conducted in the Sahel region of Africa. In this section of Africa, summertime high intensity rainfall results in sand-topped soil crusts that are subsequently wind eroded in the windy winter months. Frontal storms from the east and Harmattan winds originating in the Sahara to the north deposit more than 2 Mg ha1 yr1 of nutrient-rich dust on this region (Drees et al., 1993). The landscape is characterized by plateaus of skeletal, acidic soil interspersed with valleys of sandy soil used to grow millet. Where vegetation is present, increased deposition of the nutrient-rich dust results in up to 20 cm of relatively neutral, humus rich soil (Chappell, 1996; Chappell et al., 1996). Chappell et al. (1996) endeavored to elucidate the factors controlling soil redistribution by relating topographic attributes, soil properties, and vegetative cover. An unvegetated site in the valley floor was identified and, by fitting an exponential model to the 137 Cs profiles in that location, a reference inventory of 2066 ± 125 Bq m2 was determined (Chappell, 1995). Among the findings was that 137Cs concentrations of soil from the plateau and from trapped aeolian dust was ten times greater than the sandy soils of the valley. It was concluded that water erosion was more important on the plateau than in the valley and that the presence of vegetation, more than any other factor, controlled the patterns of soil redistribution by wind (Chappell et al., 1996; Chappell, 1996). Chappell et al. (1998a) estimated that net loss was only about 50–60% of the total erosion due to the vegetation trapping of wind-entrained sediments. Due to the complex pattern of soil redistribution in this landscape, it has been concluded that actual rates may be as much as four times larger than original estimates made by previous investigators (Chappell et al., 1998b). Chappell et al. (1998a,b) used the mass balance model of Zhang et al.

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

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(1990) for wind eroded source areas and a modified proportional model for areas in which deposition was dominant. The use of these two models for aeolian redistribution studies has been validated using detailed, multi-year field measurements (Van Pelt et al., 2007). The estimates of soil redistribution rates in this area of Africa have been further refined by using geostatistical techniques coupled with remote sensing to enhance the spatial detail of sparse 137 Cs measurements (Chappell, 1998). It has also been shown from data obtained at this site that nested sampling of 137Cs activities and the application of geostatistics results in better estimates of soil redistribution than single measurements made along a transect (Chappell, 1999). He further noted that using available geomorphological information could be used to limit the number of sampling necessary to obtain good estimates. This same nested sampling pattern was utilized in a subsequent study of soil redistribution in a wind erosion affected area of England where the estimate was again improved by the use of geostatistical techniques that are more efficient than a more extensive and expensive gridded sampling approach (Chappell and Warren, 2003). Although there was very evident burial of field margins by aeolian sediment, the authors could find little evidence of net soil loss from the region. Farmers of the Sahel were approached with soil loss rates from their valley farm fields of 30 Mg ha1 yr1 as estimated using the 137 Cs technique (Warren et al., 2003). The farmers were less concerned about the long-term effects of soil loss than they were about the short-term loss of productivity caused by wind erosion effects such as stand blowout or burial and loss of nutrients. The authors state that, due to the depth of the sandy soil, erosion may continue for several years before the fields lose their ability to produce crops and government imposed conservation programs may threaten community stability. 7.5. Asian investigations There has been great interest in wind erosion processes in P.R. China due to the degradation of large areas of western and northern China from wind erosion. In the far western region of China is the Qinghai-Tibetan Plateau, a high altitude arid area that is prone to wind erosion. Yan et al. (2001) sampled multiple landforms and land use areas in the north-central part of the plateau and in the southern part of this vast region. They report 137Cs reference inventories of 2376 and 982 Bq m2 for the north-central and southern locations, respectively. By using the proportional model, they determined soil loss rates of 84.1, 69.4, 30.7, and 21.8 Mg ha1 yr1 for shrub-stabilized coppice dunes, semi-stabilized dunes, dryland farm fields, and grasslands, respectively, and for the entire Qinghai-Tibet Plateau, they estimated an annual soil loss rate of 47.6 Mg ha1. Their estimated erosion rates for the individual management areas agree with a detailed, spatially intensive wind erosion modeling investigation for China and fall into their ranges of slight to highly intensified and the integrated mean value is in the moderate range. Deposition of aeolian sediments has also been documented in lakes of this region (Yan et al., 2002). In a later investigation, several sites in three high altitude locations were sampled for 137Cs inventories in the southern and far western portion of the Tibet Plateau (Zhang et al., 2007). The authors reported 623 Bq m2 as the reference 137Cs inventory for the arid far western location and 1065 and 1740 Bq m2 for the two southern sites. The larger reference inventory of the two southern locations was from the higher altitude of the two locations. They found that alpine meadows have the lowest wind erosion potential but that soil erodibility increased markedly with disturbance from livestock grazing and woodcutting. They found semi-stabilized aeolian sand and mobile sand to have the greatest

erosion rates. They concluded that restricting livestock grazing, woodcutting, and excessive grassland cultivation are the keys to controlling wind erosion in this region. In north-central China, northeast of the Qinghai-Tibetan Plateau is an area that is a complex of many landforms and land uses including mountains, dune fields of many morphologies, sandy grasslands, sandy shrublands, and farm fields. In a 6 km2 area of this landscape, 125 points were sampled for 137Cs inventories and the results kriged to interpolate and prepare contour plots based on 7350 points (Zhang et al., 2003). The authors report a 137 Cs reference inventory of 2319 Bq m2, similar to the reference activity found for the northern area of the Qinghai-Tibetan Plateau (Yan et al., 2001). They found that wind erosion on farmland were much greater than grassland which had a slightly greater erosion rate than stabilized dunes. Of the 125 points sampled, 110 showed net loss of 137Cs and soil and only 15 showed accretion of 137Cs and sediment. These few depositional areas were primarily in shifting sand lands and on the toes of shrub-stabilized dunes, especially at the lee toes of the dunes. Li et al. (2005) used the 137Cs technique to study wind and water erosion processes in the Wind-Water Erosion Crisscross Region of the Loess Plateau that is eroded by primarily wind in the winter and spring and by primarily water in the summer and autumn. Reference 137Cs inventories are reported as 1580 Bq m2. They sampled the 137Cs inventories on a cultivated rounded hill with equal slopes using two transects, east to west and north to south, that crossed at the summit. The erosive winds of the winter season come primarily from the northeast, and were assumed to only erode the north and east aspects of the hill. The difference between erosion rates on these aspects, for which 137Cs inventories indicated greater overall erosion, and the erosion on the lee slopes, the researchers estimated that 30% of the sediment contributed to the Kuye River was mobilized by wind and 70% was mobilized by water. West of the Qinghai-Tibetan Plateau, Muminov et al. (2010) used 137Cs, lithogenic radioisotopes, and cosmogenic radioisotopes to study processes and rates of soil redistribution along a transect from the base of the mountains to the center of a closed basin in Uzbekistan. They noted that the least disturbed areas had depths of 137Cs transport and inventories that were fairly consistent and might represent good candidates for reference areas. 137Cs inventories for these locations ranged from 1900 to 2700 Bq m2, values consistent with those noted for some locations on the Qinghai-Tibetan Plateau. They concluded that water was the primary erosive force on sloping land along the rock outcrop, steep pediment, and alluvial apron near the base of the mountains, especially where livestock trails ran perpendicular to the slope contour. Along the Bajada with lesser slopes, wind erosion plays a greater role in soil redistribution due to the heavy use by livestock, especially near water sources, pens, and fence lines. The higher activity of 137Cs on unwashed plant material collected at the end of the dry windy season, compared with plant material that was washed or collected at the end of the growing season before wind had entrained disturbed soils tends to substantiate the seasonal dominance of wind erosion in this region. More recently, Funk et al. (2011) studied multiple processes in wind erosion, transport, and deposition for grazing lands and cultivated fields using the 137Cs technique at a site in north-central China near Inner Mongolia. Using the flow considerations of Goossens (1988, 1996, 2006) to locate a reference location in this highly disturbed environment, they obtained a 137Cs reference inventory of 1967 ± 102 Bq m2. They reported that the concentration of 137 Cs in suspended sediment tends to increase with height above the surface, that the increase appeared to be a increase greatly at heights of 0.1 m and above from the surface, and that there was strong deposition in vegetated locations downwind of eroding

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

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fields. The further noted that the 137Cs activity of the upper 2 cm of soil in these depositional areas was approximately equal to that observed in the suspended sediment. The authors demonstrated a threshold vegetation height between erosional and depositional environments of 5 cm. They also stated that, based on meteorological records, erosion was probably greater in the past than currently measured. In a depositional environment downwind of an eroding source area, Chen et al. (2009) have used the 137Cs pulse in 1963 from stratospheric fallout, a peak in 1972 from local atmospheric testing in the Lop Nor region of Northwest China, and a 1986 peak from the Chernobyl accident and fire to look at temporal patterns of erosion in northern China. They assumed the pulse depths coincided with surface incorporation and used a postulated Soil Wind Erosion Index (SWEI) determined by the ratio of coarse sand >0.25 mm to fine sand <0.1 mm in the downwind deposition area to determine the relative intensities of wind erosion events between the pulses. The presence of the 1986 pulse from Chernobyl, however, is problematic for most erosion researchers using the 137 Cs technique. 8. Non-stratospheric fallout sources of anthropogenic radioisotopes The use of 137Cs and the radioisotopic technique to soil redistribution rates depends on the assumption of temporally predictable fallout from the stratospheric circulation during the period during and following atmospheric testing of nuclear weapons. When sources other than this stratospheric fallout contribute to the 137 Cs inventory in the soil serious errors in rate estimation will occur and the method is invalid. There are other documented sources of radioisotopic contamination at local and regional scales. In this section, I will review these sources and the areas in which they have limited the practical use of 137Cs inventories to estimate soil redistribution rates. 8.1. Low yield and failed detonations Soil-borne anthropogenic radionuclides come from sources other than globally distributed stratospheric fallout. The site of the first atmospheric detonation is still radioactive from many elements encased in the vitrified sands including fission daughters 90 Sr and 137Cs as well as unspent 239Pu and 240Pu from the bomb’s nuclear fuel which is <3% 240Pu (Parekh et al., 2006). Likewise, soils and sediments in areas downwind from nuclear test sites also contain anthropogenic radioisotopic signatures from the local detonations and failed tests. Increased activity of radioisotopes in the soil has been documented for much of the area surrounding the USA’s Nevada Test Site (NTS) including areas in Utah (Beck and Krey, 1983). Cizdziel et al. (1998) studied the attic dust in locations surrounding the NTS and found that 137Cs/239+240Pu ratios of the sediments were smaller than expected for global fallout, 34 ± 4, indicating a significant contribution from incomplete fission of 239 Pu and 240Pu in failed tests. They also state that redistribution of soil-derived dust from highly contaminated NTS soil is evident in the region. 8.2. Accidental releases from peaceful nuclear activity Anthropogenic radioisotope presence in soils may also be from non-weapon sources. 238Pu used in the power supply of orbiting satellite SNAP-9A was released when it burned up on re-entry to Earth’s atmosphere in April, 1964 (Cigna et al., 1987), spreading nearly a kg of 238Pu across all the continents. Areas near industrial facilities that process radioisotopes are often contaminated from

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accidental releases. Sites such as Hanford, Washington and Rocky Flats, Colorado in the USA, a nuclear reprocessing facility and a nuclear materials production facility, respectively, have isotopic ratios different from global fallout, indicating local sources of contamination from accidental release (Hodge et al., 1996; Litaor and Ibrahim, 1996). Contaminated soils at the Rocky Flats site have spread by wind erosion and deposition, impacting previously uncontaminated soils to the east of the site. Soils of the southern Ural Mountains in the former Soviet Union are contaminated with radioisotopes (Oughton et al., 2000; Skipperud et al., 2000) accidentally released from the Mayak reprocessing facility in 1957 and from wind-blown contamination originating at a near-by nuclear waste disposal site (Beasley et al., 1998). Many of the better known industrial releases are related to accidental releases from nuclear power plants. In March, 1979 radioactive 131I was released from the Three Mile Island nuclear generating station in Pennsylvania, USA (Gerusky, 1981). A few years later, a core meltdown and fire at the Chernobyl power station in the Ukraine contaminated many Eurasian soils with 137Cs (Mabit et al., 2008) including sites near Chernobyl (Muramatsu et al., 2000), in England (Walling and Quine, 1991; Michel et al., 2001), in Austria (Bossew et al., 2007), in northwest Syria (Al-Masri, 2006), and further east in China (Yan et al., 2002). Using the measured 240Pu/239Pu ratio of 0.408, Muramatsu et al. (2000) determined that the source of contamination was from local sources rather than global fallout. In the early years after the Chernobyl accident, the presence of 134Cs allowed quantification of the 137 Cs in the soil that was from Chernobyl (Mitchell et al., 1990). However 134Cs has a half-life of 2.1 years and thus is currently present at levels below detection (Bunzl et al., 1994) making the correction impossible. More recently, the nuclear power station at Fukushima Daiichi on the east coast of Japan was damaged in a catastrophic earthquake and tsunami and emitted 137Cs at levels nearing those observed at Chernobyl. These emissions have been transported and deposited in Japan and areas across the Pacific Basin (MacKenzie, 2011) as well as areas of central Asia (Bolsunovsky and Dementyev, in press). These post-fallout pulses of 137Cs have caused and will continue to cause difficulty for soil erosion researchers in affected locations. 9. Conclusions and future prospects Since the discovery that anthropogenic radioisotopes are adsorbed to and move with soil particles, many researchers have used and misused this technique, most notably with 137Cs, to estimate rates of soil erosion and deposition. The advantages of using 137 Cs are many, they outweigh the limitations, and using 137Cs to estimate historic rates of wind erosion and subsequent deposition is a valid technique provided that researchers take due diligence to use the technique in a responsible manner. One misuse of 137Cs has often been the result of the time and cost of gamma spectroscopy and the economy of using a small number of samples to estimate the redistribution rate over a large and sometimes complex landscape. Spatial variability of soil properties such as texture, organic matter content, and runoff and infiltration influence the initial inventory of 137Cs in an undisturbed landscape and often affect the erodibility of the soil in the landscape of interest. The use of spatially intensive sampling, including nested sampling, and the use of geostatistics and associated modeling approaches such as the Kalman filter promise to improve the accuracy of redistribution estimates made using 137Cs. Another misuse of the technique is the continued use of the simple proportional model when other, more rigorous models have been developed and validated. Although detractors of this technique have voiced their concerns and doubts, they have failed to point to an alternative method of estimating multi-decadal erosion rates in a cost effective manner.

Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004

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The use of 137Cs inventories to perform local calibrations of predictive models for soil erosion has been underutilized and represents a valuable use of the technique. Unfortunately, the decay of 137 Cs will rapidly make the use of it as a tracer extremely difficult and of questionable accuracy. At present, only about a third of the original 137Cs from global fallout exists and with each decade, local and regional contamination from industrial sources is masking the contribution from global fallout. 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Please cite this article in press as: Van Pelt, R.S. Use of anthropogenic radioisotopes to estimate rates of soil redistribution by wind I: Historic use of 137Cs. Aeolian Research (2013), http://dx.doi.org/10.1016/j.aeolia.2012.11.004