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a U.S. Geological Survey, 345 Middlefield Rd. MS 465, Menlo Park, CA 94025, and Dep. of Plant and Soil Sci., Univ. of Delaware, Newark, DE 19717
b Dep. of Earth and Planetary Sciences, The Johns Hopkins Univ., Baltimore, MD 21218
c Dep. of Plant and Soil Sciences, Univ. of Delaware, Newark, DE 19717
* Corresponding author (yarai{at}usgs.gov)
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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6.0), but P adsorption gradually decreased with decreasing pH near the soil pHwater (5.0). The overall findings suggest that P losses from the can be suppressed by an increase in the P retention capacity of soils via (i) an increase in the number of lime applications to maintain soil pHwater at near targeted pH values, and/or (ii) alum/iron sulfate amendments to provide additional Al- and Fe-based adsorbents.
Abbreviations: CEC, cation exchange capacity • DE, Delaware • Delmarva, Delaware–Maryland–Virginia Peninsula • EDS, energy dispersive spectrometer • EMP, electron microprobe • Ev, Evesboro • ICP–AES, inductively coupled plasma–atomic emission spectroscopy • OM, organic matter • Os, Osier • PL, poultry litter • Pm, Pocomoke • PZSE, point of zero salt effect • Sf, Sassafras • WDS, wavelength dispersive spectrometer • XRD, x-ray diffraction
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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To fulfill the answers to the questions above, the objectives of this study were to investigate P adsorption capacity, P desorbability, and its relation to operationally defined P fractions using the following comprehensive laboratory experiments. First, the PL-amended DE soils were well characterized using indirect and direct approaches (e.g., sequential inorganic P fractionation and EMP analyses). Second, the relationships between desorbable P fractions and operationally defined P fractions defined as major chemical components in the Mehlich 3 chemical extraction (e.g., NH4 oxalate and NH4F) in the P-rich DE soils were investigated by combining short- and long-term desorption experiments and chemical extractions. While short-term (24-h) P desorption experiments were conducted on the soils that have been pretreated with specific chemical reagent treatments, long-term (30-d) P desorption experiments were performed without any chemical pretreatment. Third, to better predict the P partioning processes in the soils, the effect of pH (4.5–7.5) and initial P concentrations (0.1–1 mM) on P retention capacity were investigated in the PL-amended DE soils using batch adsorption experiments.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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Mineralogical Analyses
Organic matter and metal oxides were removed from the soils using sodium hypochlorite and citrate–bicarbonate–dithionite methods, and then sand, silt and clay fractions were obtained using centrifugation-sedimentation methods described by Jackson (1956). Freeze-dried samples were finely ground using a mortar and pestle for further use. Powder XRD measurements were made to determine the bulk mineralogy. The data were collected from 5 to 95° 2
with a Philips x-ray powder diffractometer (graphite monochromatized Cu K
radiation, 0.05° 2 step size, and 4 s count time per step). All samples were analyzed as random mounts using the back-packed procedure. The detection limit for crystalline phases by bulk XRD analyses is approximately 2% by weight.
Inorganic Phosphorus Fractionation
Bulk soil samples were ground using a mortar and pestle before chemical extractions. The inorganic P fractionation method described by Kuo (1996) was slightly modified to optimize the extraction processes by removing Ca from limed agricultural soils.
Inorganic P fractions were divided into soluble P, P in the amorphous Al-Fe oxide fraction, P in crystalline Al oxides and/or varicsite (AlPO4) like phases, and P in crystalline Fe oxide and/or strengite-like (FePO4) and/or hydroxyl apatite-like phases. During the extraction processes, all tubes were shaken at 250 rpm on an end-over shaker. After the extraction, the samples were passed through 0.45-µm membrane filters before total P analyses using ICP–AES. Since most of the extracts could contain >1 mM of inorganic or organic ligands that interfere with the original ammonium molybdenum analyses (He et al., 1998), ICP–AES for total P was chosen over a molybdenum reactive P analysis. Between each extraction step, samples were washed with 50 mL of 0.1 M NaCl solution three times, and centrifuged at 268 g for 5 min to recover the paste for the next extraction steps.
Figure 2 shows the sequential inorganic P fractionation Steps (a–e). In Step (a), soluble P fractions were extracted using 0.7 g of oven-dried soils (<2 mm) that were suspended in 50 mL of 1 M NH4Cl solutions for 30 min. Paste samples were then pretreated with 1.0 M ammonium acetate (pH 5.5) for 1 h to remove exchangeable Ca that could readily precipitate out as Ca oxalate during the oxalate extraction. In Step (b), P adsorbed on amorphous Fe and Al oxides and/or amorphous Al, Fe-P components (e.g., variscite and strengite) were fractionated after Step (a) using the ammonium oxalate extraction method (Loeppert and Inskeep, 1996). In Step (c), P sorbed on crystalline aluminum oxides and/or AlPO4 fractions was extracted using 50 mL of 0.5 M NH4F (pH 8.2) solution for 1 h. In Step (d), P in crystalline iron oxides and/or strengite-like phases were extracted after Steps (a) and (b) using the dithionite-citrate-bicarbonate method. In Step (e), P-Ca fractions (e.g., hydroxyl apatite phases) were extracted in 50 mL of 0.25 M H2SO4 for 1 h after which the samples were treated in Steps (a, b and d).
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30 µm thick) mounted on quartz slides (2.54-cm diam. and 1.6 mm thickness) using methyl 2-cyanoacrylate adhesive. Selected materials and media were chosen to minimize trace metal or metalloid contamination. Analyses were performed on the thin sections using an electron microscope operating at 15 kV potential and a 30 nA beam current. Eight 128 by 128 pixel x-ray maps (pixel size: 0.91 mm) were obtained with a 1-s dwell covering 117 by 117 µm2. Elemental analyses for Si and Fe were collected using a Be window energy dispersive spectrometer (EDS), while Al, S, Ca, and P were collected with wavelength dispersive spectrometers (WDSs). Maps were obtained at a magnification of x750 to avoid WDS defocusing. The EDS x-ray spectra were also collected at points of interest.
Phosphorus Adsorption Experiments
Phosphate retention capacity has been often evaluated at a fixed pH value (e.g., pHwater) and an initial phosphate concentration. However, such results are often difficult to interpret in predicting the fate and transport of P in agricultural soils since (i) soil pH values fluctuate before and after liming, and (ii) dissolved P concentrations in the surface soils can vary before and after PL amendments at current or reduced application rates. For these particular DE soils, recommended target pH values are generally 6.0 for Ev, Sf, and Os soils and 5.6 for the Pm soil (Sims and Gartley, 1996). To better understand P adsorption capacities on the long-term PL-amended soils, we investigated the P adsorption behavior as a function of pH (3.0–8.0) and initial P concentrations (0–1 mM). It is difficult to select an ideal dissolved P concentration range in the laboratory experiments to evaluate the P adsorption capacity in PL-amended soils since PL is randomly broadcast on agricultural fields (i.e., nonuniform applications). We suspect the dissolved [P] is not uniformly distributed throughout the field. Instead of estimating the ideal dissolved [P] range based on the PL application rate, the upper [P] limit (i.e., 1 mM 31 mg L–1) in this experiment was chosen based on the [P] range that was reported in laboratory studies. Shreve et al. (1996) previously reported that dissolved P concentrations from PL-amended loamy soils, which were amended at a rate of 0.5 g of PL in 50 g of soils, decreases from 25 to 7 mg L–1 at pH 5.0 with increasing time (0–300 d) (Shreve et al., 1996). Since the application rate that was used in Shreve's study is slightly higher than the typical PL application rate on DE soils (i.e., 0.2 g of PL in 50-g soils; equivalent to 9 Mg ha–1), the reported upper limit of dissolved [P] (31 mg L–1) in our experiment should be sufficient to simulate the field conditions in DE soils.
Air-dried soil samples (i.e., 1.2 g of the amended Ev, Os, Pm, and Sf) and 24 mL of 0.1 M NaCl were shaken at 100 rpm on an orbital shaker for 20 h at 25 ± 2°C. Each soil sample was adjusted to the field pHwater using 0.1 M HCl and NaOH. An appropriate amount of the NaH2PO4 stock solutions (i.e., 1 and 5 mM in 0.1 M NaCl at pH 5.0) was added to ensure initial P concentrations of 1 mM, 0.5 mM, and 0.1 mM. The pH was readjusted every 6 to 12 h by adding 0.1 M HCl or 0.1 M NaOH. After 48 h, the soil suspensions were centrifuged at 478 g and passed through a 0.45-µm filter. The modified ammonium molybdenum method (He et al., 1998) was used to analyze concentrations of dissolved orthophosphate. The same method was also used to analyze concentrations of dissolved orthophosphate in the following the desorption experiments. Total adsorbed P (mg kg–1 of soil) was estimated based on differences between [Pinitial] and [Premaining] with respect to the amount of soils (in kilograms) used in the experiments.
Long-term Phosphorus Desorption Experiments
A batch replenishment method was used to investigate long-term (30-d) P desorption. Air-dried soils (1.2 g) were placed in 50-mL high-speed polycarbonate centrifuge tubes. Thirty milliliters of 0.1 M NaCl solutions were added that were adjusted to the soil pHwater values. The soil suspensions were shaken on an end-over-end shaker at 200 rpm. Every 24 h, the tubes were centrifuged at 11950 g for 5 min, and the supernatants were replaced with the same P-free 0.1 M NaCl solutions. This process was repeated 30 times (for a 30-d period).
In addition to the Mo active P analyses, the filtrate was also analyzed for total Fe and Al using ICP–AES.
Short-Term Phosphorus Desorption Experiments
A stirred-flow method was chosen for the short-term desorption study. In this technique, the adsorbents were exposed to a greater mass of ions than in a static batch system, and the flowing solution continuously removed reaction products (desorbed and detached species) (Sparks, 1989). Air-dried soils (0.6 g) were placed on a stirred flow chamber assembled with a 0.45-µm membrane filter (Sparks, 1989), and 7.6 mL of 0.1 M NaCl solution at the soil pH were added. The soil suspensions were preequilibrated for 24 h at 100 rpm before the desorption experiments. The influent solution was pumped at a flow rate of 0.3 mL min–1 into the reaction chamber that was mixed at 300 rpm. Nine-milliliter effluents were collected every 30 min using a fraction collector. Filtrates were analyzed for dissolved P as described earlier. The desorption experiment was also repeated on the same soil samples which had received three different sequential chemical extractions, namely NH4Cl, NH4Cl + ammonium oxalate, and NH4Cl + NH4F as described in the inorganic P fractionation methodology section. Oxalate and fluoride ligands represent major chemical components (e.g., ammonium oxalate and ammonium fluoride) in the Mehlich 3 chemical extraction. The NH4 oxalate and F extractions were performed after the NH4–acetate pretreatment to remove exchangeable Ca, since oxalate and F are known to precipitate out with exchangeable Ca in soils (Smillie and Syers, 1972). While the NH4F extraction was performed after the NH4Cl + NH4 oxalate extractions in the previous inorganic P fractionation analyses to remove only operationally defined crystalline Al-P fractions (e.g., crystalline variscite and P sorbed crystalline aluminum oxides), the NH4F extraction in these experiments was performed right after the NH4Cl treatment to remove both amorphous and crystalline Al-P fractions in soils. After these chemical treatments, the soils were washed twice with 30 mL of 0.1 M NaCl solutions adjusted to the soil pH, and freeze-dried for the desorption experiments. The combined techniques (i.e., chemical treatment and desorption) were employed to investigate the relationships between operationally defined inorganic P fractions and the desorbable P fractions.
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RESULTS AND DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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Inorganic Phosphorus Fractionation
The results of the inorganic P fractionation of the PL-amended soils are summarized in Table 3. The total inorganic P values for each soil (i.e., summation of each inorganic P fraction measured using the modified Mo blue method) were slightly less than the total P values in Table 2, but greater than the total inorganic P values that were estimated by the method described by Bowman (1989). This is probably because of the use of latter ICP in the analyses. This highlights the importance of the inorganic P fraction in the long-term PL-amended soils. The predominance of inorganic P fractions in the long-term PL-amended soils was previously observed by other researchers (Sharpley and Smith, 1995).
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The chloride anion extractable P (i.e., weakly held P) ranged from 22 to 49 mg kg–1 after the long-term amendments in all soils. Although these values are small with respect to total inorganic P, this increased chloride anion extractable P fraction could have a significant impact on eutrophication since the typical eutrophication inducible P level is as low as 0.04 mg L–1 (Sims et al., 1996). Several field studies have indicated a strong correlation between water extractable soluble P and rainfall simulated dissolved P or runoff (Pote et al., 1999; Sharpley and Moyer, 2000).
While the ammonium oxalate extractable P comprises >60% of the total P, the Ca-P fractions remained consistently low (<5% of total inorganic P). Although the Ca concentrations in PL are generally high (i.e., >15000 mg kg–1)(Arai et al., 2003), the acidic soil pH might have suppressed the formation of alkaline earth metal phosphate precipitates (e.g., apatite) that are generally thermodynamically favored at neutral to alkaline pH.
Electron Microprobe Analyses
Electron microprobe images of Al, Si, S, P, Ca, and Fe in the Ev and Pm soils are shown in Fig. 3a and Fig. 3b, respectively. In the low-OM Ev soil (Fig. 3a), P is generally associated with Al and Fe (Regions 1 and 2 indicated by arrows), and to a much lesser extent with Ca (as indicated by a Region 3 circle). In the high-OM Pm soils (Fig. 3b), Fe and S are associated with P. While P and S oxyanion adsorption on Fe based adsorbents (e.g., iron oxide and pyrite) can be speculated, the P-S association might also indicate P retention on S-rich soil humic components. Phosphorus is known to strongly react or associate with organic carbon components in soils (Bloom, 1981; Hens and Mercks, 2001; McDowell and Sharpley, 2001). As observed in the Ev soil, Ca is poorly correlated with P in the Pm soil (as indicated by Region 5). To semiquantitatively investigate the relative concentration of elements, energy dispersive x-ray spectra were taken at the P hot spots (indicated as Regions 2 and 4 in Fig. 3a and Fig. 3b, respectively). The most intense peak was from Si followed by Al, K, P, and Fe peaks were also observed (Fig. 4a and Fig. 4b). These Al and Si peaks might be attributed to a mixture of quartz and albite that were previously confirmed via XRD analysis (Table 1) and/or other aluminosilicate minerals such as kaolinite and illite. Although anions (e.g., arsenate) are known to weakly adsorb on quartz at acidic pH (Xu et al., 1988), large quantities of quartz and albite might retain a substantial amount of P overall in these sandy soils. The strong P association with Al and Fe probably represents P adsorption complexes on Al-Fe oxides and/or Al-Fe-P bulk and coprecipitates and/or chelated complexes on metal-humic components. In Fig. 4a and Fig. 4b, intensity of P peaks at the P hot spot is much stronger in the Ev soils than in the Pm soils, suggesting that the P speciation in the low- and high-OM soils are different. If one assumes that bulk and coprecipitation mechanisms are predominant in these high P (>800 mg kg–1 of total P) soils, a possible explanation for the different P localization in low- and high-OM soil matrices would be the organic acid suppressed amorphous Fe- and Al-P precipitation mechanisms. Several researchers have reported the organic acid-suppressed Al- and Fe-P precipitation reactions (Struthers and Sieling, 1950; Swenson et al., 1949). Whereas P is readily precipitated out in low-OM soils, the precipitation reactions were suppressed in high OM, resulting in more evenly distributed adsorbed species in soil matrices. Since the release of P is highly influenced by the solid-state P speciation, P precipitates in low-OM soils might be more susceptible than sorption complexes in high-OM soils on changes in the P concentration gradient in soil solutions. Overall, the EMP analyses agree with the results of inorganic P fractionation that P is generally associated with Al and Fe and to a lesser extent with Ca.
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There was also a trend in all soils for overall adsorption to increase with increasing [P]o at pH values of 3.0 to 8.0 (Fig. 5a–5d). Interestingly, net positive adsorption was only observed at pH 4.2 to 6.2 in the Ev soil that was spiked with 1 mM P, and at 4.2 to 7.3 in the Os soil that was spiked with 
0.5 mM P. When the Sf and Pm soils were spiked with [P]o > 0.1 mM, the soils exhibited net positive adsorption at pH values 4.8 to 7.0 and 4.2 to 7.0, respectively. Unlike the Ev soil, the Sf, Pm, and Os soils showed a net positive adsorption capacity at similar pH ranges even when the soils were spiked with [Po] = 0.5 mM. A possible explanation for the positive adsorption at these pH ranges could be the formation of amorphous Mg/Ca-PO4 precipitates, which are generally thermodynamically stable at pH > 5.5. However, one can expect that those amorphous Ca-PO4 minerals rapidly undergo dissolution processes when the soils are buffered back to an acidic soil pHwater (5). The previous chemical extraction and EMP analyses indicate the absence of Ca-PO4 minerals in these soils. At moderately acidic pH values (<5.0), most of the soils (except for the Os soil) did not show any significant adsorption capacity—even the soils that were spiked with high [Po] (i.e., 1 mM). This suggests that the P sinks in these soils have nearly reached their maximum P sorption capacities. The average net negative surface charge density at pHsolution > PZSE of the soils (3.08–3.67, Table 2) might be repulsing the negatively charged P solution species (i.e., H2PO4–). However, the actual P release mechanisms are difficult to postulate based on these macroscopic data alone since several minerals simultaneously undergo dissolution at acidic pH.
Long-term Phosphate Desorption Experiments
Results of the long-term P desorption are shown in Fig. 6a. After 30 replenishment processes, total desorbable P was highest in the Pm which had the lowest total inorganic P (Table 3), followed by Ev, Sf, and Os. The degree of P desorbablity is not correlated with differences in the amount of inorganic P, suggesting a complex desorption mechanism in these soils. It seems that the extent of P desorption in these soils are different. Phosphate desorption in the low-OM soils (e.g., Ev and Sf) decreases after 25 replenishments (i.e., time); however, there were steady and continuous desorption reactions in the higher-OM soils (e.g., Pm) at >25 replenishments. Slow P desorption behavior from the inorganic soil components (i.e., ferrihydrite, goethite, and kaolinite) have been extensively reported by many researchers (Bar-Yosef et al., 1988; Madrid and Posner, 1979; Ryden et al., 1977). These results indirectly suggest that these mineral components might retain the labile P fractions in low-OM soils like Ev and Sf. However, there is no straightforward explanation to support the continuous P desorption reactions observed in the higher-OM soils.
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100 mg kg–1 in the Sf soil, total Al release was as high as 400 mg kg–1 in the Pm soils, suggesting that each soil has different P release mechanisms associated with Al and Fe. There are some consistencies in the molar ratios of total desorbed P with respect to the Fe and Al release after 30 d. The P/Fe and P/Al ratios are 7.33 and 0.94 in Ev, 4.32 and 1.00 in Sf, 4.54 and 0.68 in Os, and 8.78 and 0.66 in Pm, respectively. While the P/Al ratios remain 
1 in all soils, the P/Fe ratios are consistently higher (i.e., >4) than the P/Al ratios, indirectly suggesting that the major labile P fractions might be associated with Fe more than Al. Several researchers have reported a similar relationship; that is, labile P in runoff is more correlated with oxalate extractable Feox than Alox (Schroeder et al., 2004). The previous XRD analyses did not detect any Al-P and Fe-P precipitates in these soils. This suggests that the simultaneous P, Al, and Fe release might be attributed to (i) dissolution of P sorbed on the Fe and Al based or complexed adsorbents (e.g., metal oxides) and/or (ii) desorption of ternary Al/Fe-PO4 complexes on soil components. Several indirect macroscopic studies have postulated the formation of mixed metal (Al, Cu, and Zn)-P complexes on soils and soil components (Bleam et al., 1991; Bolland et al., 1977; Clark and McBride, 1985).
Short-Term Phosphate Desorption Experiments
The continuous-flow method was also used to measure the extent of P desorption reactions in the well-drained soils (e.g., Ev and Sf). Results of the short-term desorption are presented in Fig. 7a to 7d. The short-term desorption experiments were conducted not only on the PL-amended soils, but also on chemically pretreated PL-amended soils. The chemical treatments were NH4Cl, NH4Cl + NH4F, and NH4Cl + NH4 oxalate extractions that are previously described in the inorganic P fractionation section. The desorption curves after each treatment show different P desorbability from remaining P fractions.
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24 h) was observed in the untreated soils followed by the NH4Cl treated, NH4Cl + NH4F treated, and NH4Cl + NH4 oxalate treated soils. This indirectly suggests that the operationally defined chloride, fluoride, and oxalate extractable P fractions might be controlling the P release in these soils. Although the poorly drained Os and Pm soils released >200 mg kg–1 of total inorganic P, <170 mg kg–1 was released from the well-drained Ev and Sf soils. As we observed in the EMP analyses, different solid-state P species are likely to be present in the well-drained Ev soil and the poorly drained Pm soil. Different solid-state P species might be a limiting factor causing the different P release behavior in these soils.
The NH4Cl treatment resulted in a small reduction in total desorbable P in most of the soils (i.e., Ev, Os, and Pm soils). It seems to appear that a significant reduction (i.e., 40% of total desorbable P of the untreated sample) was occurring in the Sf soil (Fig. 7b); however, the reduction in the total desorbable P is not significantly different from the other soils when we compare the mass basis values of the reduction in total desorbable P after NH4Cl treatment in Fig. 7 and the NH4Cl extractable P in Table 3 for each soil. The reduction in total desorbable P after NH4Cl treatment is approximately 30, 65, 50, and 15 mg kg–1 for the Ev, Sf, Os, and Pm soils, respectively; whereas for the NH4Cl exctractable P in Table 3, total desorbable P is about 44, 49, 22, and 25 for Ev, Sf, Os, and Pm soils, respectively. This suggests that remaining P species after the NH4Cl extraction in all soils behaved in a similar fashion.
The NH4Cl + NH4F treatment resulted in the second highest P release from all soils. While it resulted in 50% reduction in total desorbable P in the low-OM Ev and Sf soils, it had the most significant impact (i.e., >73% reduction) on the poorly drained and high-OM Os and Pm soils, suggesting that the role of Al with respect to the P release is more important in high-OM soils (Os and Pm soils) than low-OM soils (Ev and Sf soils). This result possibly infers the presence of labile OM-Al-P complexes in high-OM soils as previously reported by some researchers (Hens and Mercks, 2001; McDowell and Sharpley, 2001). Furthermore, fluoride extractable pools indicate both crystalline and amorphous Al-P materials in the soils since the NH4F chemical treatment was employed before the oxalate treatment. If one assumes the fluoride ions were effectively extracting these Al-P fractions, the reduction in the P release via the fluoride treatment indicates that total desorbable P in these soils is controlled by Al. The simultaneous release of Al was also observed in the previous long-term P desorption experiments. Maguire and co-workers have previously reported a similar correlation between Al-associated P and the bioavailable P in biosolid-amended Atlantic Coastal Plain soils (Maguire et al., 2000).
We have reported the results for the important labile Fe-P fractions in the previous long-term desorption experiments, and the results from the short-term desorption study seem contradictory. We suspect that the different findings (i.e., labile Fe-P vs. Al-P fraction) could be due to differences in the desorption experimental techniques. Whereas the batch replenishment technique in the long-term desorption experiments simulates poorly drained conditions, possible readsorption reactions of desorbed species can occur in the closed reaction vessels. The continuous flow-through system that was used in the short-term desorption experiments simulates well-drained conditions, resulting in minimal readsorption processes during the experiment. It may be possible that the different drainage properties of these soils might cause different P release mechanisms that are associated with Fe and Al.
Interestingly, the desorption results from the fluoride treated soils show a small quantity (52–60 mg kg–1 of total desorbable P at 57 chamber volumes) of desorbable P in all soils. Since the previous XRD data showed no clear evidence for the presence of crystalline Fe-P, Al-P, and Ca-P, it is possible that crystalline phases contribute little to P desorption. Rather, the amorphous Fe phases (i.e., oxalate extractable Fe) control the P desorption after the fluoride treatment. When the total desorbable P after the fluoride treatment was compared in low-OM Ev and Sf soils, this suspected amorphous Fe-P fraction seems to release more P in the Sf soil than in the Ev soil. The previous inorganic fractionation data showing that Feox was greater in the Sf (590 mg kg–1) than the Ev soil (353 mg kg–1) (Table 3) supports the hypothesis that bioavailable amorphous Fe-P fractions are greater in the Sf than in the Ev soil.
Similarly, when the total desorbable P after the fluoride treatment was compared in the high-OM Os and Pm soils, the effect of fluoride on total desorbable P cannot be explained by the differences in the remaining amorphous Fe-P content. On the basis of Feox concentrations in the inorganic fractionation data (Table 3) that the Pm soil has less Feox content than in the Os soil (337 and 752 mg kg–1, respectively) (Table 3), the Pm soils should release less P from the amorphous Fe-P fraction that is left after the F treatment. However, this is not the case. The total desorbable P values after F treatment in the Os and Pm soils are almost equal (i.e., 60 mg kg–1) after 57 chamber volumes (Fig. 7c and Fig. 6d). The Pm soil is unexpectedly releasing more P even though the Feox content is almost half of the Os soils (Table 3). This may be because the Pm soil is more poorly drained (thus its much greater OM in Table 1), resulting in more periodic saturated conditions than found in the Os soils. The reduced conditions in the Pm soils might have promoted greater P association with amorphous Fe minerals that are newly formed under the simultaneous reductive dissolution and reprecipitation reactions during the reduction or reoxidation reaction. Vadas and Sims (1999) previously suggested similar P retention mechanisms to explain enhanced P retention capacity on the A horizon of Pm soils after reduction or reoxidation reactions (Vadas and Sims, 1999).
Implications for Selecting the Best Nutrient Management Plans for the Phosphorus-Rich Delaware Soils
In these P-rich soils, we found that P was still strongly retained near targeted pH values of lime (i.e., 6.0 for Ev, Sf, and Os soils and 5.6 for the Pm soil), however, P adsorption gradually decreased with decreasing pH near soil pHwater (5). As a result of long-term PL amendments, there was no evidence, based on XRD analyses, for the presence of crystalline Al and Fe phosphate precipitates and apatite phases. Instead, the inorganic P fractionation analyses showed high levels of oxalate extractable P, Al, and Fe fractions (P = 615–858 mg kg–1; Al = 1215–1478 mg kg–1; Fe = 337–752 mg kg–1) that are susceptible to slow or continuous P desorption at moderately acidic soil pHwater along with soluble P fractions (i.e., operationally defined chloride extractable P). Various soil P tests [e.g., water soluble P, ammonium oxalate extraction based extractions (Mehlich 1 and 3)] have been often used to predict the desorbable P fractions from PL-amended P-rich soils. Although these soil P tests were originally developed for agronomic purposes, the results were often positively correlated to dissolved P in runoff and leachates (Pote et al., 1996). These indirect evidences, which were based on simple batch extractions and linear regressions analyses, suggest that soil test P might potentially be utilized to predict the labile P (Chardon and van Faassen, 1999; Pautler and Sims, 2000; Sims et al., 2002; van der Zee and van Riemsdijk, 1988). Our study results provide additional credence evidence to the previous findings.
On the basis of our research findings, several remediation or nutrient management methods on surface soils can be considered or reevaluated for providing the most environmentally feasible and effective ways to reduce the labile P concentrations from the P-rich soils. These include (i) maintaining soil pHwater near neutral, and (ii) having effective adsorbents (e.g., amorphous Al and Fe minerals) that can help in retaining soluble P to suppress the P desorption processes.
Perhaps the best way to minimizing future losses of P from high P Delmarva soils is to stop amending them with PL. However, this is not a reasonable solution in view of the large quantity of PL that must be disposed of. Additional and realistic remediation or nutrient management methods are needed to reduce soluble P in the immediate future. If PL application continues at the current or reduced rates, it seems that alkaline characteristics of PL would potentially neutralize the soil pH; however, the soil pH is unlikely to remain at near targeted pH of limed soils because (i) randomly broadcasted PL will not uniformly raise pH throughout the entire agricultural field, and (ii) soil pH strongly buffered back to moderately acidic pH. An increase in the number of lime application would be an environmentally feasible choice to raise soil pHwater values without increasing P input in soils only if it does not facilitate the macronutrient (e.g., Mn and Zn) deficiency for crops. One can expect that an increased lime application would effectively maintain soil pH at near neutral in low-OM soils (e.g., Ev and Sf soils); however, it seems less effective for high-OM soils (i.e., Pm and Os soils) since weakly acidic soil OM strongly buffered at moderately acidic pH. As the EMP analyses indicated, different P solid-state speciation were already present in low- and high-OM soils (e.g., possibly precipitates vs. adsorption complexes, respectively). Accounting for the different P speciation (i.e., different release mechanisms) in the high and low-OM soils, additional nutrient management practice (e.g., chemical amendments) should be recommended in addition to an increase in lime.
Chemical amendments {e.g., alum [Al2(SO4)3] and iron sulfate} can provide additional adsorbents for soluble P in soils (e.g., amorphous Al/Fe minerals), resulting in suppressed P release. Both of these chemical amendments have been shown to effectively reduce the soluble P runoff from PL-amended Arkansas soils (Shreve et al., 1995; Shreve et al., 1996). Alum amendment on PL has been currently evaluated as a potential remediation or nutrient management method to reduce the soluble P in the Atlantic Coastal Plain soils. The recent studies indicated that alum-amended PL on soils has some potential in reducing the bioavailable P in acidic sandy DE soils (Sims and Luka-McCaffery, 2002; Staats et al., 2004) and P adsorption complexes on amorphous Al-oxides in the alum-amended PL are likely to be controlling the P release (Peak et al., 2002). Addition of alum to moderately acidic Atlantic Coastal Plain soils will needs further investigation to determined the long-term stability of P and the impact of soil acidity resulting from the Al inputs. On the contrarily, the iron sulfate amendment might be used in DE sandy soils without jeopardizing future soil acidity in soils. Assuming that amorphous Fe oxyhydroxide is controlling the P release as results of iron sulfate amendments, the reductive dissolution of Fe-oxides under seasonally reduced conditions could possibly induce the P release. Field and laboratory scale studies must be conducted to evaluate the redox sensitivity of the FeSO4/PL-amended DE soils with respect to P release.
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ACKNOWLEDGMENTS |
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Received for publication June 30, 2004.
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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