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TECHNICAL REPORTS

Heavy Metals in the Environment

Major and Trace Elements of Selected Pedons in the USA

USDA Natural Resources Conservation Service, Soil Survey Laboratory, 100 Centennial Mall North, Room 152, Mail Stop 41, Lincoln, NE 68508-3866

^* Corresponding author (rebecca.burt{at}nssc.nrcs.usda.gov).

Received for publication February 4, 2003.

	ABSTRACT

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Few studies of soil geochemistry over large geographic areas exist, especially studies encompassing data from major pedogenic horizons that evaluate both native concentrations of elements and anthropogenically contaminated soils. In this study, pedons (n = 486) were analyzed for trace (Cd, Co, Cr, Cu, Hg, Mn, Ni, Pb, Zn) and major (Al, Ca, Fe, K, Mg, Na, P, Si, Ti, Zr) elements, as well as other soil properties. The objectives were to (i) determine the concentration range of selected elements in a variety of U.S. soils with and without known anthropogenic additions, (ii) illustrate the association of elemental source and content by assessing trace elemental content for several selected pedons, and (iii) evaluate relationships among and between elements and other soil properties. Trace element concentrations in the non-anthropogenic dataset (NAD) were in the order Mn > (Zn, Cr, Ni, Cu) > (Pb, Co) > (Cd, Hg), with greatest mean total concentrations for the Andisol order. Geometric means by horizon indicate that trace elements are concentrated in surface and/or B horizons over C horizons. Median values for trace elements are significantly higher in surface horizons of the anthropogenic dataset (AD) over the NAD. Total Al, Fe, cation exchange capacity (CEC), organic C, pH, and clay exhibit significant correlations (0.56, 0.74, 0.50, 0.31, 0.16, and 0.30, respectively) with total trace element concentrations of all horizons of the NAD. Manganese shows the best inter-element correlation (0.33) with these associated total concentrations. Total Fe has one of the strongest relationships, explaining 55 and 30% of the variation in total trace element concentrations for all horizons in the NAD and AD, respectively.

Abbreviations: AD, anthropogenic dataset • AM, arithmetic mean • ASD, arithmetic standard deviation • CEC, cation exchange capacity • CEC₇, cation exchange capacity determined by NH₄OAc at pH 7 • GM, geometric mean • GSD, geometric standard deviation • MD, median • NAD, non-anthropogenic dataset

	INTRODUCTION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
ELEMENTAL ANALYSIS of the fine-earth (<2 mm) or other particle-size fractions has been used in the study of soil properties on a pedon, landscape, or ecosystem basis by providing information on geologic processes and parent material uniformity (Chapman and Horn, 1968; Kaup and Carter, 1987); pedogenesis (Brimhall et al., 1991; Wilson et al., 1996); or mineral weathering, composition, and phase quantification (Nettleton et al., 1970; Dubbin et al., 1993; Amonette and Sanders, 1994). Mineral weathering during pedogenesis results in translocation and/or accumulation of major or trace elements via chemical and biological processes such as leaching, podzolization, and oxidation–reduction (Davies, 1980; Kabata-Pendias and Pendias, 1992; Mausbach and Richardson, 1994; Pierzynski et al., 1994; Wilson et al., 1996). Determination of soil processes and resulting pedogenic changes often involves elemental evaluations of horizons within the solum, between the solum and parent material, and along bioclimatic gradients. Pedogenic evaluations using elemental analysis include the silica to sesquioxide ratio as a soil weathering index (Jackson, 1979); ratios of alkali cations (e.g., Ca and K) and the relatively stable elements Zr and Ti as measures of pedogenic changes and geological uniformity (Sudom and St. Arnaud, 1971; Muhs et al., 2001); total and fractionated forms of P as indices of soil development (Walker and Syers, 1976; Tiessen et al., 1984; Singleton and Lavkulich, 1987; Cross and Schlesinger, 1995; Burt and Alexander, 1996); and the relative accumulation of Fe and Al and depletion of P as a function of climate indicative of pedogenic gradients (Tedrow, 1968; Jenny, 1980; Bockheim, 1980; Birkeland et al., 1989). Geochemical studies have also used trace elements to document soil processes such as element release and transport (weathering and profile leaching) in Spodosols (Jersak et al., 1997) and element immobilization (oxidation–reduction and associated pH changes) in wetlands (Gambrell, 1994). Elemental data are also useful information in environmental studies of soil and water, for example, metal contamination through atmospheric deposition (Storm et al., 1994; Burt et al., 2000) or transport in surface or ground waters (Mesuere et al., 1991; Jones et al., 1997; Martens and Suarez, 1997; Whatmuff, 2002).

Quantifying the total or total extractable pool of trace elements has been a common procedure for studies assessing environmental levels or background amounts of trace elements in soils (Shacklette and Boerngen, 1984; Shuman, 1985; Holmgren et al., 1993; Mermut et al., 1996; Chen et al., 1999). This pool is routinely determined not only because these data are important to any overall assessment of the fate, bioavailability, and transport of trace elements (Shacklette and Boerngen, 1984; Holmgren et al., 1993; Wilson et al., 2001) but also due to the lack of a widely applicable method to assess the bioavailable fraction.

Many studies have examined relationships among elements (major and trace) and between elemental concentrations and other soil properties in noncontaminated soils. Aluminum and Fe were strongly related to total P, with Al (r² = 0.83) explaining most of the variation in 22 important U.S. soils (Burt et al., 2002a). Total Al and Fe explained 29% of the variation in baseline concentrations of trace elements in Florida surface soils (Chen et al., 1999) and total Al strongly correlated with selected trace elements (Pb, Cd, and Cr with r² = 0.88, 0.81, and 0.45, respectively) in Louisiana coastal wetlands (Pardue et al., 1992). Trace element concentrations have also been related to clay and CEC in 40 mineral soils in Florida (Ma et al., 1997) and to physical surface area (e.g., clay, organic matter, CEC) in the southeastern USA (Shuman, 1985). Other studies have related elemental extractability (speciation) or potential reactivity of trace elements to other soil properties such as total concentrations and/or soil type, pH, texture, organic matter, and carbonates (Chlopecka et al., 1996; Ma and Rao, 1997; Kabala and Singh, 2001; Andersen et al., 2002).

Previous studies developing correlations between trace elements and other soil properties have generally focused on non-anthropogenic sources of trace elements or have not evaluated native and anthropogenic sources separately. Elements derived from anthropogenic sources probably have correlations that differ with other soil properties. For example, a recent study (Andersen et al., 2002) used differences in correlation to speculate on possible anthropogenic additions.

This study was initiated to quantify a wide variety of soils (with and without known anthropogenic additions) from across the USA. The objectives are to (i) determine ranges of concentrations of selected elements in selected pedons from across the USA and soils with known anthropogenic additions, (ii) discuss trace element concentration for several selected pedons to illustrate the association of elemental source and content, and (iii) evaluate relationships among and between elements and other soil properties as factors governing amounts and distribution of elements.

	MATERIALS AND METHODS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Study Soils
Pedons from across the coterminous USA (Fig. 1) as well as Hawaii, Alaska, Virgin Islands, Guam, and Puerto Rico were analyzed for major and trace elements. Selected major horizons were analyzed for each of the 486 pedons, with surface horizons defined here as any surface mineral material (e.g., A and E horizons) and underlying materials as B and C horizons (Soil Survey Staff, 1999a). In addition, satellite samples were analyzed for several selected pedons. Satellites are defined as pedons that are generally sampled several meters from the primary pedon, providing information on spatial variability of soil properties. Included within this total dataset are soils with anthropogenic inputs (AD) as well as soils regarded as having no known anthropogenic additions (NAD).

View larger version (28K): [in this window] [in a new window]   Fig. 1. Map illustrating locations of sampling sites in the study. Each location is representative of multiple samples (e.g., horizons and/or satellites). Clusters of points in a particular geographic location represent an area of intensive study.

Field and Laboratory Methods
Pedons were described by standard soil survey methods (Soil Survey Staff, 1962; Soil Survey Division Staff, 1993) and sampled using procedures described in Soil Survey Investigations Report 42 (Soil Conservation Service, 1984; Soil Survey Staff, 1996). Pedons were classified by U.S. soil taxonomy (Soil Survey Staff, 1999a).

Samples were analyzed by procedures described in Soil Survey Investigations Report 42 (Soil Survey Staff, 1996). Alphanumeric codes in parentheses beside each method represent specific analytical procedures. Bulk samples were air-dried and sieved in the laboratory to remove rock fragments that were >2 mm (1B1). All standard analyses were performed on air-dried <2-mm soil with resulting data reported on an oven-dry basis.

Samples (about 35 g of <2-mm soil) for elemental analysis were ground in a silicon nitride ball mill to <150 µm. Trace elements (Cd, Co, Cr, Cu, Hg, Mn, Ni, Pb, Zn) were determined on soil samples (0.5 g, <150 µm) digested in a covered vessel with 9 mL HNO₃ + 3 mL HCl (aqua regia) in a microwave (180°C for 7 min). Digestate was diluted to a final volume of 50 mL. Extracts were analyzed by inductively coupled plasma atomic emission spectrometry (ICP–AES), axial mode and ultrasonic nebulization with internal standardization and inter-elemental correction. Total Hg was determined by cold-vapor atomic absorption. Total trace element concentrations reported herein, unless otherwise specified, refer to the sum of Cd, Co, Cr, Cu, Mn, Ni, Pb, Zn, and Hg. Extracts for major elements (Al, Ca, Fe, K, Mg, Na, P, Si, Ti, Zr) were prepared by microwave digestion (180°C for 9.5 min) using 0.25 g soil (<150 µm) combined with 4 mL HF + 9 mL HNO₃ + 3 mL HCl. Following digestion, boric acid was added to neutralize HF (2.5% in 100 mL final volume) and extracts analyzed by ICP–AES, radial mode and cross-flow nebulizer. This digestion is considered a more rigorous digest than by aqua regia alone, as HF is noted for the efficient dissolution of silicate minerals and is often used in conjunction with aqua regia to ensure the digestion of organic constituents of soils (Sawhney and Stilwell, 1994; Wilson et al., 1997a). Quality assurance samples (blank, duplicate, and Certified Reference Material, CRM) were included for every 28 samples in the digestion process. Our elemental recoveries and relative standard deviations (RSD) for the CRM (Loam Soil C; High-Purity Standards, Charleston, SC), based on similar digestion procedures, ranged from 90 to 115% and 11%, respectively.

Particle-size analysis was determined by sieve and pipette (3A1), following pretreatments to remove organic matter and soluble salts, and chemical dispersion with sodium hexametaphosphate. For statistical data evaluation, clay was estimated for samples with poor dispersion (1500 kPa water retention; clay percentage of >0.6) (Nettleton and Brasher, 1983; Soil Survey Staff, 1999a):

Organic C (6A1c) was determined by acid-dichromate digestion with FeSO₄ titration and total C and N by dry combustion (6A2f and 6B4b, respectively). Soil pH (H₂O) was determined on a 1:1 soil and water paste (8C1f). Exchangeable cations (Ca, Mg, K, Na) were extracted with buffered (pH 7.0) NH₄OAc and measured by atomic absorption spectrometry (AAS). The cation exchange capacity (CEC₇) was determined by Kjeldahl titration of the NH₄–saturated soil (5A8b). Acidity was extracted with BaCl₂–TEA (pH 8.2) (6H5a). Base saturation by CEC₇ was determined by dividing the sum of NH₄OAc-extractable bases by CEC₇ and multiplying by 100 (5C1). Exchangeable Al was extracted with 1 M KCl on samples with pH (H₂O) of <5.5 and determined by ICP–AES (6G9b). Bray P-1 (6S3e) was determined colorimetrically with a spectrophotometer. The percent CaCO₃ was determined by treating samples with HCl, with evolved CO₂ measured manometrically. Dithionite–citrate (6C2b, 6G7a, 6D2a) extracts were analyzed for Fe, Al, and Mn by AAS. Acid-oxalate (6C9a, 612a, 6S8a, 6V2b) extracts were analyzed for Fe, Al, P, and Si by ICP–AES. Optical grain counts were conducted on the coarse-silt (0.02–0.05 mm), very-fine sand (0.05–0.10 mm), or fine-sand (0.10–0.25 mm) fraction (7B1a). Crystalline clay minerals (<2 µm) were identified by X-ray diffraction (XRD) analysis (7A2i).

Statistical Methods
Characterization data for individual pedons are available online at http://soils.usda.gov (USDA, 2003). Trace element data will soon be available at this USDA website as well. Soils were analyzed using routine statistical analysis (Manugistics, 1998), including arithmetic mean and standard deviation (AM and ASD, respectively), geometric mean and standard deviation (GM and GSD, respectively), median (MD), range, analysis of variance (ANOVA), correlation, and simple and multiple linear regression. Medians and GMs were compared using the Kruskal–Wallis and t test procedures, respectively, at the 95% confidence level (p < 0.05). Selection of the statistical tool for comparison (e.g., GM or MD) was based on applicability of test to dataset(s). Independent variables were determined for simple and multiple linear regression models at the 90, 95, and 99% confidence levels (p < 0.1, 0.05, and 0.01, respectively). Before regression models were developed, data sets were normalized by log base 10 transformations to meet the criteria of normality. Regression coefficients (coefficients of determination) and correlation coefficients are herein designated as r² and r, respectively.

	RESULTS AND DISCUSSION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Total Dataset: Trace and Major Element Concentrations
The relative abundance of major elements (Table 1; n = 1903) as indicated by AMs is as follows: Si > Al > Fe (AM = 285, 47, and 30 g/kg, respectively), with lesser amounts of K, Ca, Na, and Mg. The trace elements in greatest abundance are Mn and Zn (AM = 508.9 and 95.2 mg/kg, respectively), with Cd and Hg having the lowest soil concentrations (AM = 0.32 mg/kg and 188 µg/kg, respectively). These relative trends among elements are also found for the MDs and GMs and are in good agreement with other studies of soils in the USA (Chen et al., 1999; Holmgren et al., 1993; Gough et al., 1988; Shacklette and Boerngen, 1984). For most elements, AM > MD > GM, with the MDs typically closer in value to the lower range of elemental concentration than to the upper range and closer to the GMs than to the AMs, indicating positive skewness of data. The ASDs for most major and trace elements are either equal to or greater than the AMs for these elements, indicative of the wide variability in these soils and their significant departures from normal distributions. The AM and ASD are considered good estimates of the geochemical abundance of an element, whereas the GM and GSD are considered better estimators of the maximum likelihood for most geochemical data (Chen et al., 1999; Gough et al., 1988).

The ranking of GMs for total trace elements by Order (Table 2; n = 1020) is as follows: Andisol > Oxisol > (Inceptisol, Mollisol) > (Aridisol, Alfisol) > Vertisol > (Spodosol, Entisol) > Ultisol. This high amount of trace elements in Andisols (1900 mg/kg) may be related to the greater sequestering ability of organic compounds and noncrystalline oxides and hydroxides found in soils of the Andisol order. The presence of noncrystalline Fe oxides within soils has been related to the decreased mobility of some trace elements (Kuo et al., 1983), due in part to the higher specific surface area (Davis and Leckie, 1978; Nalovic, 1978) compared with crystalline Fe oxides. In this study, Andisols have one of the greater Fe and Al pools, with GM = 70 and 76 g/kg, respectively.

A second aspect related to soil orders is presentation of these data on a weight percentage rather than volume percentage basis. Volumetric conversion incorporates bulk density and particle density of individual horizons. Expressing data on a volume basis improves the quantification of elemental gains and losses during pedogenesis (Brimhall et al., 1991; Chadwick et al., 1990). It should be recognized that the lower density of Andisols relative to other orders would result in a relative decrease in trace elements on a volume basis in these soils compared with other orders. In this study, data are expressed on a weight basis because (i) bulk density was not available for all soils and (ii) particle density varies widely depending on parent material and mineralogy and was not measured. In environmental research, trace elemental concentrations are typically expressed on a weight basis, and it is especially suitable for a study with a large, diverse geographic area.

A commonly used weathering index is the silica to sesquioxide ratio (Jackson, 1979). Oxisols have one of the lowest ratios (1.6) across all horizons, with the largest ratio found in Entisols (12.5). Oxisols are highly weathered soils with subsurface horizons usually low in CEC and nearly devoid of weatherable minerals, and the clay fraction is composed primarily of low-activity minerals (e.g., kaolinite), with lesser amounts of gibbsite, goethite, and hematite (Lynn et al., 2002). Entisols have diverse mineralogies (reflective of parent materials) and typically have little evidence of weathering (Lynn et al., 2002). In this study, Entisols have a relatively high GM Si (338 g/kg) compared with other Orders (nearly twice the concentration in Andisols, 119 g/kg), due in part to the inclusion of the Psamments (sandy Entisols) (Soil Survey Staff, 1999a).

Non-Anthropogenic Soils: Example Pedons Illustrating Trace and Selected Major Element Concentrations
The variability of elemental concentration within or between soil orders can best be illustrated by evaluating data from individual pedons (Table 3). Magmatic source materials, from which many Andisols develop, have distinctly different geochemistries. The trace element composition of an Andisol from Hawaii (formed from a highly weathered basaltic lava) is much higher relative to the other pedons. The Ap1 horizon of the Hilo (medial over hydrous, ferrihydritic, isohyperthermic Acrudoxic Hydrudand) pedon (Soil Survey Staff, 1999a) has between 3 and 26 times higher concentration of Co, Cr, Cu, and Ni relative to the A horizon of the Threebear (medial over loamy, amorphic over mixed, superactive, frigid Alfic Udivitrand) pedon, a soil developed from andesitic ash from Idaho. The trace element content of the Threebear, Mexico (fine, smectitic mesic Vertic Epiaqualf), and Estelle (medial over loamy-skeletal, ferrihydritic over mixed, superactive Andic Haplocryod) pedons is relatively similar. The upper horizons of the Estelle and Mexico pedons are developed from loess from different sources, but generally have similar amounts of trace elements. Largest differences exist between Pb and Hg, with higher amounts of these elements in the Mexico and Estelle pedons, respectively.

Non-Anthropogenic Soils: Trace and Selected Major Element Concentrations by Soil Horizon
The GMs by horizon (Table 4) indicate that trace element concentrations generally appear to be greater in the A and/or B than in C horizons, suggesting accumulation via pedogenetic or possible anthropogenic inputs of unknown origin on surface horizons. Cadmium appears evenly distributed with depth in profiles of NAD soils, ranging from 0.15 to 0.17 mg/kg. Similar results were found by Ma et al. (1997) in a study of 40 mineral soils in Florida, with Cd distribution generally independent of Orders (Spodosols, Ultisols, Entisols) and fairly constant in the soil profile. However, the lack of discernible depth trends for Cd may also be related in part to the low concentrations typically found in soils, often in amounts below instrumental detection limits (<5% of all samples in study). Mercury shows higher values in surface horizons compared with B and C horizons (78 versus 45 and 53 µg/kg, respectively). These results may be due in part to the greater likelihood of organic substances in surface materials, which can serve as strong complexing agents for Hg (Inacio et al., 1998).

The Dia (fine-loamy over sandy or sandy-skeletal, mixed, superactive, mesic Oxyaquic Haploxeroll) pedon (Table 6) is derived from alluvium of mixed origin (granitic sources from the Sierra Nevada plus andesite, rhyolite, tuff, and basalt from local ranges of Tertiary Age). Major elements for horizons composing the Dia pedon range from 9 to 13% Al₂O₃, 2 to 5% Fe₂O₃, and 51 to 52% SiO₂, similar in composition to basalts found in the western USA (e.g., Table Mountain, Colorado; Lassen Peak, Madera, and Mount Shasta, California; and Colfax County, New Mexico) (Clarke, 1924). Dia has elevated amounts of Hg (798 µg/kg) compared with other pedons in Table 6, increasing with depth to 1956 µg/kg in the C2 horizon (46- to 69-cm depth, data not shown) but decreasing below the lithological discontinuity to 1257 µg/kg in the 2C4 horizon (91- to 152-cm depth, data not shown). The Hg depth function in this pedon suggests that the Hg is probably from native origins. In general, non-anthropogenic Hg rocks are often found in association with igneous geologic materials, particularly vein deposits with gold and silver, and often deposited in association with sulfide minerals such as pyrite and marcasite in rocks such as shales. It is found in a few locations concentrated in the form of a sulfide mineral, cinnabar. Many of the chief deposits of mercurial ores are in the neighborhood of igneous rocks from which they were probably derived. Deposition of cinnabar was observed in a geothermal system near Steamboat Springs, Nevada (Clarke, 1924).

The ID-34 (loamy-skeletal, siliceous, subactive, shallow Typic Dystrocrept) pedon (Table 6), located near the Norris Geyser Basin in Yellowstone National Park (Wyoming), is developed from hydrothermally altered rhyolitic, coarse-textured glacial till. In Yellowstone, acid sulfate areas develop in volcanic areas of topographically elevated landscapes or areas with relatively impermeable, underlying geologic units. The alteration or acid sulfate weathering process involves H₂SO₄ produced from exsolution of gaseous H₂S gas from magmatically heated ground water. This gas is oxidized to H₂SO₄ in the vadose zone by both inorganic reactions and bacteria. The influx of gas and hydrothermal fluids in this soil accelerate the weathering process on minerals resulting in the concentration of certain elements (e.g., 27.8 mg/kg Pb, 300 mg/kg Hg) and depletion of others (36.9 mg/kg Zn). These acidic conditions solubilize Fe and Al, creating a whitish soil matrix with low extractable cations and base saturation. This process is similar to the pedogenetic process that occurs in albic horizon formation (Soil Survey Staff, 1999a).

The Plymouth (mesic, coated Typic Quartzipsamment) pedon (Table 6) from New York City was sampled as part of a study to provide quantitative data for mapping units and determine limitations for urban land use. The surface layer of Plymouth has 621.2, 119.7, 64.0, and 0.91 mg/kg Pb, Zn, Cu, and Cd, respectively. This pedon exemplifies the results of an investigation of 98 pedons in this survey area showing trace element concentrations in surface materials (<20 to 30 cm deep) as Pb > Zn > Cu > Cr > Co > Cd, with concentrations of Pb, Zn, and Cd elevated several-fold relative to underlying horizons. These results suggest that concentrations occurring at these lower depths may be useful as reference or "baseline" data, depending on the history of site and uniformity of the soil parent material. Elevations in concentration were also found in the current (year 2000) levels of Pb, Zn, and Cd compared with 10 years ago at the same locations (e.g., 138.2 versus 118.1 mg/kg Pb, respectively), suggesting sources of soil Pb other than automotive Pb and/or possible redistribution of Pb through human activities. This study also found some variation in total metal concentrations in surface horizons in several New York City parks (e.g., with upper ranges of 683.8 mg/kg Pb in Central Park versus 221.3 and 234.2 mg/kg in Van Cortland and Pelham Parks, respectively), but limited spatial variability in satellites short distances from the primary pedon regardless of park location (e.g., 376.9 to 470.1 mg/kg Pb in a Central Park pedon and 125.8 to 200 mg Pb in a Van Cortland Park pedon; data not shown). Similar elevated elemental concentrations were found in an urban study of Glebe, Australia (Markus and McBratney, 1996) as well as significant variation in total metal concentrations due to soil disturbance and distance from the street (e.g., explaining 30% of the variation in Pb levels). In the New York City study, variation between and within parks was attributed to distance from the street, differences in fill, and possible dilution effect of sampling protocol (e.g., 0- to 5-cm depth versus 0- to 20-cm depth [entire horizon]).

The Beaverell (loamy-skeletal over sandy or sandy-skeletal, mixed, superactive, frigid Aridic Argiustoll) pedon (Table 6) is a contaminated soil resulting from long-term copper smelting in the area of Anaconda and Deer Lodge Valley, Montana. This pedon exemplifies other contaminated soils in this area, with elevated trace element concentrations as follows: Cu > As > Zn > Pb (Burt et al., 2000). In addition, concentrations are greatest in surface horizons, decreasing abruptly with depth in total amounts and reactive forms (Burt et al., 2003). In general, there is decreasing soil contamination with distance from the smelter but unlike New York City park pedons, spatial variability of total trace element concentrations in satellites is relatively high (e.g., 966.3 versus 1236.3 mg/kg), attributed to microerosion and deposition processes (from wind or water) around vegetation that accumulates more of the metals than surrounding eroded and barren soils. Similar results were found by Dudka et al. (1995), relating significant spatial variability in trace element concentrations to the effect of main point sources of contaminants, namely smelters. Even though Beaverell is <1 km from the old smelter stack, some trace elements in the New York City Plymouth pedon (e.g., 621.2 mg/kg Pb) are more elevated than in Beaverell (474.5 mg/kg Pb). Trace elements from anthropogenic sources have been found to be more mobile than those of native origin (Karczewska, 1996; Ma and Uren, 1998; Wilcke et al., 1998). Type of anthropogenic source can also affect the fate and transport of trace elements in soils. For example, oxidized Pb compounds emitted from active smelters have been shown to be more soluble than solid phases of Pb in mining soils, and as such are considered more bioavailable (Davis et al., 1994; USEPA, 1990). This may have implications for the bioavailable fraction of Pb and possibly other contaminants in New York City and similarly contaminated urban soils, as most particles emitted from power plants, incinerators, and automobiles are in the smaller particle-size fraction (e.g., <2 µm in diameter) and tend to be dispersed farther and have higher total and soluble metal contents (Tiller, 1989; Hughes et al., 1980; Page et al., 1979; Nriagu, 1984).

The Brazilton (fine, mixed, active, nonacid, thermic Alfic Udarent) pedon (Table 6) is a reclaimed surface mine soil in Crawford County, Kansas. Total trace elements in surface materials of the Brazilton are nearly one-half as much as found in one of the most contaminated sites (Beaverell pedon) in the Deer Lodge Valley study, but the Brazilton has greater amounts of Co, Cr, Mn, and Ni than the Beaverell pedon, with 22.2, 72.6, 1885.4, and 56.0 mg/kg, respectively. Unlike other pedons with anthropogenic inputs presented here, the Brazilton pedon has been altered anthropogenically by placement of soil and rock material naturally enriched in certain trace elements higher in the soil profile (zone of active soil weathering) than found in the originally formed soil. Soil development appears to be initially quite rapid in these reclaimed Kansas soils probably due to placement of unweathered, soluble rock materials (shales) into the upper layers of soil material. The Brazilton pedon is located on an overgrazed fescue (Festuca spp.) pasture, suggesting a potential for erodibility. This pedon has a surface pH of 4.7, 62% base saturation, and 34 g/kg dithionite–citrate-extractable Fe, with total C and N levels (23.9 and 2.49 g/kg, respectively) reflective not only of the biologically active pool but also of the organic C and N associated with carbonaceous shale, and as such is less available for chemical complexation and retention of trace elements. These results have implications for the greater likelihood of release and transport of some trace elements from this soil into the environment compared with native soils.

The Genola (fine-silty, mixed, mesic Typic Calcixerept) pedon (Table 6) from San Pete County, Utah was sampled as part of a study to provide baseline data to help quantify the likelihood of environmental problems resulting from manure and fertilizer application of P to soils in this area. The site was one of a several areas selected across Utah for comparing native soil P levels with those sites that had received large manure additions of P. The Genola pedon was developed in a mixed sandstone–shale alluvium, with a surface pH of 8.2 having 290 g/kg CaCO₃ and similar amounts of carbonates throughout the profile. While there is no measurable increase in total C from the application of manure, total P is elevated in the upper 7 cm of the manured Genola (1452 mg/kg P) versus the native Genola (920 mg/kg, data not shown), with nearly comparable P levels below the 7-cm depth (753 and 832 mg/kg, respectively). Extractable Bray P-1 is also elevated at this depth in the manured Genola versus native Genola, with 23.3 and 0.9 mg/kg, respectively (data not in tables). Traditionally, Bray P-1 has been used as a plant-available or labile P index but it has also been evaluated for use in environmental studies (e.g., predictive models for P runoff; Gartley and Sims, 1994; Pote et al., 1996; Sharpley, 1996; Sims et al., 1998). Total P in the manured Genola pedon exceeds all other pedons in Table 6 with the exception of the Plymouth pedon (1577 mg/kg P). Elevated P in the manured Genola pedon is a direct result of manure application and is primarily associated with the organic fraction, whereas P in the Plymouth pedon probably resulted from multiple nonpoint sources, some of which may be in the form of insoluble lead phosphates, a proposed fate for some automotive-associated Pb in road ecosystems (Nriagu, 1984).

Anthropogenic and Non-Anthropogenic Soils: Trace Element Concentrations and Other Soil Properties
The study pedons (total dataset) vary widely in physical and chemical properties, for example, alkaline to very acid with pH (H₂O) ranging from 2.1 to 9.9; carbonates from 0 to 870 g/kg; and electrical conductivity values as high as 35.5 mS/cm. Organic constituents (C, N) are also variable, ranging from 0 to 584 and 0 to 36 g/kg, respectively. Some pedons consist of very clayey, sandy, or silty materials. The CEC ranges from 0.1 to 106.5 cmol/kg and base saturation from 1 to 100%. Extractable acidity and Al are significant components in some soils, with upper ranges of 130.5 and 24.6 cmol/kg, respectively. These upper and lower ranges of soil properties for the total dataset closely parallel those of the NAD.

Simple linear regression models were developed for total trace element concentrations of the AD and NAD (Table 7) using some of these soil properties (total Al, Fe, and Mn, CEC, organic C, pH, and clay). Models were improved when soils were divided by soil horizon in both the NAD and AD. Multiple linear regression models were also developed for total trace element concentrations in all horizons of the AD and NAD (Table 7). Relationships of total trace element concentrations with one or more soil properties provide insight into factors that govern trace element concentrations or retention in soils. These mathematical models are not intended to imply that a single physicochemical factor is responsible for these concentrations nor that the relative contribution of these factors is static, but rather varies with size and range of properties of the dataset.

View this table: [in this window] [in a new window]   Table 7. Regression and correlation coefficients of total trace element concentrations with various soil properties in non-anthropogenic and anthropogenic soils by horizon.

Total Fe has one of the strongest relationships, compared with other soil properties, for total trace element concentrations in the NAD, explaining 55, 57, 44, and 65% of the variation in all, surface, B, and C horizons, respectively. Total Fe also has the strongest relationship for total trace element concentrations in the AD, explaining 30, 33, 24, and 53% of the variation in all, surface, B, and C horizons, respectively. However, these values are relatively lower in the AD than in the NAD.

Good correlations are found between total trace element concentrations (calculated without Mn as determined by HNO₃ + HCl digest) and total Mn (as determined by HF + HNO₃ + HCl), with r values of 0.33, 0.39, 0.64, and 0.71 in all, surface, B, and C horizons, respectively, in the NAD. These results may be due in part to the high sorption capacity of Mn oxides for trace elements such as Pb, Cd, and Zn (Chao, 1972; Wilcke et al., 1998) as well as to the overall composition of the parent material.

The multiple regression model for all horizons of the NAD shows that Fe, Al, organic C, CEC, and clay can explain 68% of the variation in total trace element concentrations in these soils, whereas the model for the AD, using Fe, Al, and organic C, explains only 46% of the variation, with some of the unexplained variation in total trace element concentrations in the AD probably attributed to anthropogenic inputs.

	CONCLUSIONS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
This study evaluates trace and major elemental composition of soils across the USA, as well as the distribution of elements of major horizons within soils, including anthropogenic elemental additions. These data illustrate the influence of the important source factors in determining the amount and distribution of major and trace elements in soils—parent material, pedogenesis, and anthropogenic inputs. Without a major increase in sampling density, a continental study of trace element distribution will be unsuited to calculate background concentrations on a national or regional level. Future work to compile a dataset of trace elements in soils that adequately characterizes the distribution of trace elements across the USA will require careful evaluation of the number, soil type, and location of pedons. Our initial efforts have been directed toward determining elemental concentrations in U.S. benchmark soils to facilitate evaluation of source factors and broaden the utility of these limited data. Benchmark soils represent extensive, important, or unique soils in the USA, and therefore have been designated as having an increased level of importance relative to other soils.

This task of characterizing the geochemistry of U.S. soils is inherently difficult because they vary widely in composition not only across landscape and bioclimatic gradients, but also with depth (e.g., when related to the various dominant soil processes within the soil profile and parent material of mixed origin). Also, the separation of natural from anthropogenic sources of trace elements is difficult in soils. This study found that stronger relationships with soil properties (e.g., total Fe) occur with native rather than anthropogenic sources. Our results indicate that pedogenesis (possibly via biocycling and organic matter accumulation) increases trace element concentrations in surface versus subsurface horizons regardless of anthropogenic inputs. Determining anthropogenic inputs to soils may be facilitated by evaluating trace element distribution throughout the soil profile, with subsurface horizons possibly useful as reference or "baseline" data, depending on site history and parent material. This study also illustrates that trace element concentrations in soils from certain parent materials are naturally higher than other soils with anthropogenic inputs. Also, the task of assessing background concentration or anthropogenic additions is further complicated by the fact that these additions may be from long-range transport of contaminants of unknown sources. Nevertheless, this pursuit of soil information can be used to better define ranges of trace elements based on soil type, landscape position, or parent material. Our knowledge of the various soil processes affecting elemental amounts and distribution in U.S. soils and their relationships with other soil properties can enhance the understanding of the fate and transport of anthropogenic elements, thereby expanding the utility and application of soil survey knowledge in areas of environmental concern such as urban, mine spoil reclamation, smelter emissions, and agricultural waste applications.

	REFERENCES

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 

• Alexander, E.B. 1988. Morphology, fertility, and classification of productive soils on serpentinized peridotite in California (U.S.A.). Geoderma 41:337–351.
• Alexander, E.B., W.E. Wildman, and W.C. Lynn. 1985. Ultramafic (serpentinitic) mineralogy class. p. 135–146. In J.A. Kittrick (ed.) Mineral classification of soils. SSSA Spec. Publ. 16. SSSA, Madison, WI.
• Amonette, J.E., and R.W. Sanders. 1994. Nondestructive techniques for bulk elemental analysis. p. 1–48. In J.E. Amonette and L.W. Zelazny (ed.) Quantitative methods in soil mineralogy. SSSA, Madison, WI.
• Andersen, M.K., K. Raulund-Rasmussen, H.C.B. Hansen, and B.W. Strobel. 2002. Distribution and fractionation of heavy metals in pairs of arable and afforested soils in Denmark. Eur. J. Soil Sci. 53:491–502.
• Birkeland, P.W., R.M. Burke, and J.B. Benedict. 1989. Pedogenic gradients for iron and aluminum accumulation and phosphorus depletion in arctic and alpine soils as a function of time and climate. Quat. Res. 32:193–204.
• Bockheim, J.G. 1980. Solution and use of chronofunctions in studying soil development. Geoderma 24:71–85.[ISI]
• Brimhall, G.H., C.J. Lewis, C. Ford, J. Bratt, G. Taylor, and O. Warin. 1991. Quantitative geochemical approach to pedogenesis: Importance of parent material reduction, volumetric expansion, and eolian influx in laterization. Geoderma 51:51–91.
• Brooks, R.R. 1987. Serpentine and its vegetation: A multidisciplinary approach. Ecology, Phytogeography, and Physiology Series. Vol. 1. Dioscorides Press, Portland, OR.
• Burt, R., and E.B. Alexander. 1996. Soil development on moraines of Mendenhall Glacier, southeast Alaska. 2. Chemical transformations and soil micromorphology. Geoderma 72:19–36.
• Burt, R., M. Fillmore, M.A. Wilson, E.R. Gross, R.W. Langridge, and D.A. Lammers. 2001. Soil properties of selected pedons on ultramafic rocks in Klamath Mountains, Oregon. Commun. Soil Sci. Plant Anal. 32:2145–2175.
• Burt, R., M.D. Mays, E.C. Benham, and M.A. Wilson. 2002a. Phosphorus characterization and correlation with properties of selected benchmark soils of the U.S. Commun. Soil Sci. Plant Anal. 33:117–141.
• Burt, R., M.A. Wilson, T.J. Keck, B.D. Dougherty, D.E. Strom, and J.A. Lindahl. 2003. Trace element speciation in selected smelter-contaminated soils in Anaconda and Deer Lodge Valley, Montana. Adv. Environ. Res. 8:51–67.
• Burt, R., M.A. Wilson, M.D. Mays, T.J. Keck, M. Filmore, A.W. Rodman, E.B. Alexander, and L. Hernandez. 2000. Trace and major elemental analysis applications in the U.S. Cooperative Soil Survey Program. Commun. Soil Sci. Plant Anal. 31:1757–1771.
• Chadwick, O.A., G.H. Brimhall, and D.M. Hendricks. 1990. From a black to a grey box—A mass balance interpretation of pedogenesis. Geomorphology 3:369–390.
• Chaney, R.L., H.W. Mielke, and S.B. Sterrett. 1988. Speciation, mobility, and bioavailability of soil lead. p. 105–109. In B.E. Davies and B.G. Wixson (ed.) Lead in soil: Issue and guidelines. Science Reviews, Northwood, UK.
• Chao, T.T. 1972. Selective dissolution of manganese oxides from soils and sediments with acidified hydroxylamine hydrochloride. Soil Sci. Soc. Am. Proc. 36:764–768.
• Chapman, S.L., and M.E. Horn. 1968. Parent material uniformity and origin of silty soils in northwest Arkansas based on zirconium–titanium contents. Soil Sci. Soc. Am. Proc. 32:265–271.
• Chen, M., L.Q. Ma, and W.G. Harris. 1999. Baseline concentrations of 15 trace elements in Florida surface soils. J. Environ. Qual. 28:1173–1181.[Abstract/Free Full Text]
• Chlopecka, A., J.R. Bacon, M.J. Wilson, and J. Kay. 1996. Forms of cadmium, lead, and zinc in contaminated soils from southwest Poland. J. Environ. Qual. 25:69–79.
• Clarke, F.W. 1924. The data of geochemistry. U.S. Geol. Survey Bull. 770. U.S. Gov. Print. Office, Washington, DC.
• Cross, A.F., and W.H. Schlesinger. 1995. A literature review and evaluation of the Hedley fractionation: Applications to the biogeochemical cycle of soil phosphorus in natural ecosystems. Geoderma 64:197–214.
• Davies, B.E. 1980. Trace element pollution. p. 287–351. In B.E. Davies (ed.) Applied soil trace elements. John Wiley & Sons, New York.
• Davis, A., M.V. Ruby, and P.D. Bergstrom. 1994. Factors controlling lead bioavailability in the Butte mining district, MT, USA. Environ. Geochem. Health 16(3/4):147–157.
• Davis, J.A., and J.O. Leckie. 1978. Effect of adsorbed complexing ligands on trace metal uptake by hydrous oxides. Environ. Sci. Technol. 12:1309–1315.
• Dubbin, W.E., A.R. Mermut, and H.P.W. Rostad. 1993. Clay mineralogy of parent material derived from uppermost Cretaceous and Tertiary sedimentary rocks in southern Saskatchewan. Can. J. Soil Sci. 73:447–457.
• Dudka, S., R. Ponce-Hernandez, and T.C. Hutchinson. 1995. Current level of total element concentrations in the surface layer of Sudbury's soils. Sci. Total Environ. 162:161–171.
• Gambrell, R.P. 1994. Trace and toxic metals in wetlands—A review. J. Environ. Qual. 23:883–891.[Abstract/Free Full Text]
• Gartley, K.L., and J.T. Sims. 1994. Phosphorus soil testing: Environmental uses and implications. Commun. Soil Sci. Plant Anal. 25:1565–1582.
• Gordon, A., and C.B. Lipman. 1926. Why are serpentine and other magnesian soils infertile? Soil Sci. 22:291–302.
• Gough, L.P., G.R. Meadows, L.L. Jackson, and S. Dudka. 1989. Biogeochemistry of a highly serpentinized chromite-rich ultramafic area, Tehma, California. U.S. Geol. Survey Bull. 1901. U.S. Gov. Print. Office, Washington, DC.
• Gough, L.P., R.C. Stevenson, and H.T. Shacklette. 1988. Element concentrations in soils and other surficial materials of Alaska. U.S. Geol. Survey Professional Paper 1458. U.S. Gov. Print. Office, Washington, DC.
• Holmgren, G.G.S., M.W. Meyer, R.L. Chaney, and R.B. Daniels. 1993. Cadmium, lead, zinc, copper, and nickel in agricultural soils in the United States of America. J. Environ. Qual. 22:335–348.[Abstract/Free Full Text]
• Hughes, M.K., N.W. Lepp, and D.A. Phipps. 1980. Aerial heavy metal pollution and terrestrial ecosystems. Adv. Ecol. Res. 11:217–327.
• Inacio, M.M., V. Pereira, and M.S. Pinto. 1998. Mercury contamination in sandy soils surrounding an industrial emission source (Estarreja, Portugal). Geoderma 85:325–339.
• Jackson, M.L. 1979. Soil chemical analysis—Advanced course. 2nd ed. 11th printing. M.L. Jackson, Madison, WI.
• Jenny, H. 1980. The soil resource. Springer-Verlag, New York.
• Jersak, J., R. Amundson, and G. Brimhall, Jr. 1997. Trace metal geochemistry in Spodosols of the northeastern United States. J. Environ. Qual. 26:511–521.[Abstract/Free Full Text]
• Johnston, W.R., and J. Proctor. 1979. Ecological studies on the Lime Hill serpentine, Scotland. Trans. Proc. Bot. Soc. Edinburgh 43:145–150.
• Johnston, W.R., and J. Proctor. 1981. Growth of serpentine and non-serpentine races of Festuca rubra in solutions simulating the chemical conditions in a toxic serpentine soil. J. Ecol. 69:855–869.
• Jones, C.A., W.P. Inskeep, and D.R. Neuman. 1997. Arsenic transport in contaminated mine tailings following liming. J. Environ. Qual. 26:433–439.[Abstract/Free Full Text]
• Kabala, C., and B.R. Singh. 2001. Fractionation and mobility of copper, lead, and zinc in soil profiles in the vicinity of a copper smelter. J. Environ. Qual. 30:485–492.[Abstract/Free Full Text]
• Kabata-Pendias, A., and H. Pendias. 1992. Trace elements in soils and plants. 2nd ed. CRC Press, Boca Rotan, FL.
• Karczewska, A. 1996. Metal species distribution in top- and sub-soil in an area affected by copper smelter emissions. Appl. Geochem. 11:35–42.
• Kaup, B.S., and B.J. Carter. 1987. Determining Ti source and distribution within a Paleustalf by micromorphology, submicroscopy, and elemental analysis. Geoderma 40:141–156.
• Kuo, S.P., E. Heilman, and A.S. Baker. 1983. Distribution and forms of copper, zinc, cadmium, iron, and manganese in soils near a copper smelter. Soil Sci. 135:105–109.
• Lynn, W.C., R.J. Ahrens, and A.L. Smith. 2002. Soil minerals, their geographic distribution, and soil taxonomy. p. 691–709. In J.E. Amonette, W.F. Bleam, D.G. Schulze, and J.B. Dixon (ed.) Soil mineralogy with environmental applications. SSSA Book Ser. 7. SSSA, Madison, WI.
• Ma, L.Q., and G.N. Rao. 1997. Chemical fractionation of cadmium, copper, nickel, and zinc in contaminated soils. J. Environ. Qual. 26:259–264.[Abstract/Free Full Text]
• Ma, L.Q., F. Tan, and W.G. Harris. 1997. Concentrations and distributions of eleven metals in Florida soils. J. Environ. Qual. 26:769–775.[Abstract/Free Full Text]
• Ma, L.Q., and N.C. Uren. 1995. Application of a new fractionation scheme for heavy metals in soils. Commun. Soil Sci. Plant Anal. 26:3291–3303.
• Ma, L.Q., and N.C. Uren. 1998. Transformations of heavy metals added to soils—Application of a new sequential extraction procedure. Geoderma 84:157–168.
• Manugistics. 1998. Statgraphics Plus. Version 4.0. Manugistics, Rockville, MD.
• Markus, J.A., and A.B. McBratney. 1996. An urban soil study: Heavy metals in Glebe, Australia. Aust. J. Soil Res. 34:453–465.
• Martens, D.A., and D.L. Suarez. 1997. Selenium speciation of marine shales, alluvial soils, and evaporation basin soils of California. J. Environ. Qual. 26:424–432.[Abstract/Free Full Text]
• Mausbach, M., and J.L. Richardson. 1994. Biogeochemical processes in hydric soil formation. p. 68–127. In Current topics in wetland biogeochemistry. Vol. 1. Wetland Biogeochem. Inst., Louisiana State Univ., Baton Rouge.
• Mermut, A.R., J.C. Jain, L. Song, R. Kerrich, L. Kozak, and S. Jana. 1996. Trace element concentrations of selected soils and fertilizers in Saskatchewan, Canada. J. Environ. Qual. 25:845–853.[Abstract/Free Full Text]
• Mesuere, K., R.E. Martin, and W. Fish. 1991. Identification of copper contamination in sediments by a microscale partial extraction technique. J. Environ. Qual. 20:114–118.[Abstract/Free Full Text]
• Muhs, D.R., E.A. Bettis, III, J. Been, and J.P. McGeehin. 2001. Impact of climate and parent material on chemical weathering in loess-derived soils of the Mississippi River valley. Soil Sci. Soc. Am. J. 65:1761–1777.[Abstract/Free Full Text]
• Nalovic, L. 1978. Geochemical studies on transition elements in soils: Experimental study on the influence of trace elements on the behavior of iron and the evolution of ferric compounds in pedogenesis. Soils Fert. 41:580.
• Nettleton, W.D., and B.R. Brasher. 1983. Correlation of clay minerals and properties of soils in the western U.S. Soil Sci. Soc. Am. J. 47:1032–1036.[Abstract/Free Full Text]
• Nettleton, W.D., K.W. Flach, and R.E. Nelson. 1970. Pedogenic weathering of tonalite in southern California. Geoderma 4:387–402.
• Nriagu, J.O. 1984. Formation and stability of base metal phosphates in soils and sediments. p. 318–329. In J.O. Nriagu and P.B. Moore (ed.) Phosphate minerals. Springer-Verlag, New York.
• O'Hanley, D.S. 1996. Serpentinites: Records of tectonic and petrological history. Oxford Press, New York.
• Page, A.L., A.A. Eseewi, and I.R. Straughan. 1979. Physical and chemical properties of fly ash from coal-fired power plants with reference to environmental impacts. Residue Rev. 71:83–114.
• Pais, I., and J.B. Jones, Jr. 1997. The handbook of trace elements. St. Lucie Press, Boca Raton, FL.
• Pardue, J.H., R.D. DeLaune, and W.H. Patrick, Jr. 1992. Metal to aluminum correlation in Louisiana coastal wetlands: Identification of elevated metal concentrations. J. Environ. Qual. 21:539–545.[Abstract/Free Full Text]
• Pierzynski, G.M., J.T. Sims, and G.F. Vance. 1994. Soils and environmental quality. Lewis Publ., Boca Raton, FL.
• Pote, D.H., T.C. Daniel, A.N. Sharpley, P.A. Moore, Jr., D.R. Edwards, and D.J. Nichols. 1996. Relating extractable soil phosphorus to phosphorus losses in runoff. Soil Sci. Soc. Am. J. 60:855–859.[Abstract/Free Full Text]
• Romkens, P.F.A.M., and W. Salomons. 1998. Specific procedure for metal solid speciation in heavily polluted river sediments. Int. J. Environ. Anal. Chem. 35:89–100.
• Sawhney, B.L., and D.E. Stilwell. 1994. Dissolution and elemental analysis of minerals, soils, and environmental samples. p. 49–82. In J.E. Amonette and L.W. Zelazny (ed.) Quantitative methods in soil mineralogy. SSSA, Madison, WI.
• Shacklette, H.T., and J.G. Boerngen. 1984. Element concentrations in soils and other surficial materials of the conterminous United States. U.S. Geol. Survey Professional Paper 1270. U.S. Gov. Print. Office, Washington, DC.
• Sharpley, A.N. 1996. Availability of residual phosphorus in manured soils. Soil Sci. Soc. Am. J. 60:1459–1466.[Abstract/Free Full Text]
• Shuman, L.M. 1985. Fractionation method for soil microelements. Soil Sci. 140:11–22.
• Sims, J.T., R.R. Simard, and B.C. Joern. 1998. Phosphorus loss in agricultural drainage: Historical perspective and research. J. Environ. Qual. 27:277–293.[Abstract/Free Full Text]
• Singleton, G.A., and L.M. Lavkulich. 1987. Phosphorus transformations in a soil chronosequence, Vancouver Island, British Columbia. Can. J. Soil Sci. 67:787–793.
• Soil Conservation Service. 1984. Procedures for collecting soil samples and methods of analysis for soil survey. Soil Survey Investigations Rep. 1. U.S. Gov. Print. Office, Washington, DC.
• Soil Survey Division Staff. 1993. Soil Survey manual. USDA Agric. Handb. 18. U.S. Gov. Print. Office, Washington, DC.
• Soil Survey Staff. 1962. Soil Survey manual. USDA Agric. Handb. 18. U.S. Gov. Print. Office, Washington, DC.
• Soil Survey Staff. 1996. Soil Survey Laboratory methods manual. Version 3.0. Soil Survey Investigations Rep. 42. U.S. Gov. Print. Office, Washington, DC.
• Soil Survey Staff. 1999a. Soil taxonomy: A basic system of soil classification for making and interpreting soil surveys. 2nd ed. USDA Agric. Handb. 436. U.S. Gov. Print. Office, Washington, DC.
• Soil Survey Staff. 1999b. National soils handbook. U.S. Gov. Print. Office, Washington, DC.
• Storm, G.L., G.J. Fosmire, and E.D. Bellis. 1994. Persistence of metals in soil and selected vertebrates in the vicinity of the Palmerton zinc smelters. J. Environ. Qual. 23:508–514.[Abstract/Free Full Text]
• Sudom, M.D., and R.J. St. Arnaud. 1971. Use of quartz, zirconium, and titanium as indices in pedological studies. Can. J. Soil Sci. 51:385–396.
• Tedrow, J.C.F. 1968. Pedogenic gradients of the polar region. J. Soil Sci. 19:197–204.
• Tiessen, H., J.W. Stewart, and C.V. Cole. 1984. Pathways of phosphorus transformations in soils of differing pedogenesis. Soil Sci. Soc. Am. J. 48:853–858.[Abstract/Free Full Text]
• Tiller, K.G. 1989. Heavy metals in soils and their environmental significance. Adv. Soil Sci. 9:113–142.
• USDA. 2003. Welcome to SOILS...A site for "helping people understand soils" [Online]. Available at http://soils.usda.gov (verified 7 Aug. 2003). USDA, Washington, DC.
• USEPA. 1990. User's guide for Pb: A PC software application of the Uptake/Biokinetic Model. Version 4.0. Dep. of Res. and Development, Washington, DC.
• Walker, T.W., and J.K. Syers. 1976. The fate of phosphorus during pedogenesis. Geoderma 15:1–19.
• Whatmuff, M.S. 2002. Applying biosolids to acid soils in New South Wales: Are guideline soil metal limits from other countries appropriate? Aust. J. Soil Res. 40:1041–1056.
• Whittaker, R.H. 1960. Vegetation of Siskiyou Mountains, Oregon and California. Ecol. Monogr. 30:279–338.
• Wilcke, W., S. Muller, N. Kanchanakool, and W. Zech. 1998. Urban soil contamination in Bangkok: Heavy metal and aluminum partitioning in topsoils. Geoderma 86:211–228.
• Wild, H. 1974. Indigenous plants and chromium in Rhodesia. Kirkia 9:233–241.
• Wilson, M.A., R. Burt, W.C. Lynn, and L.C. Klameth. 1997a. Total elemental analysis digestion method evaluation on soils and clays. Commun. Soil Sci. Plant Anal. 28:407–426.
• Wilson, M.A., R. Burt, and M.D. Mays. 2001. Application of trace elements in the U.S. Cooperative Soil Survey Program. p. 324. In Proc. 6th Int. Conf. on the Biogeochemistry of Trace Elements (ICOBTE), Guelph, ON, Canada. 29 July–2 Aug. 2001.
• Wilson, M.A., R. Burt, T.M. Sobecki, R.J. Engel, and K. Hipple. 1996. Soil properties and genesis of pans in till-derived Andisols, Olympic Peninsula, Washington. Soil Sci. Soc. Am. J. 60:206–218.[Abstract/Free Full Text]
• Wilson, M.A., A.W. Rodman, G.N. White, D.P. Thoma, and H.F. Shovic. 1997b. Acid sulfate hydrothermal soil development from rhyolite flow and tuff: Yellowstone National Park, Wyoming, U.S.A. p. 219–231. In Proc. 10th Int. Working Meeting on Soil Micromorphology, Moscow. 8–13 July 1996. Printing Cervie Centre Van Gils B.V., Wageningen, the Netherlands.
• Xian, X. 1989. Effect of chemical forms of cadmium, zinc, and lead in polluted soils on their uptake by cabbage plants. Plant Soil 113:257–264.

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