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Dep. of Soil and Crop Sciences, Colorado State Univ., Fort Collins, CO, 80523-1170
* Corresponding author (jim.ippolito{at}colostate.edu)
Received for publication August 4, 2006.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Abbreviations: AB-DTPA, NH4HCO3–diethylenetriaminepentaacetic acid • L/E, Littleton and Englewood, CO wastewater treatment facility • ICP–AES, inductively coupled plasma–atomic emission spectrophotometer • Po, organic phosphorus • PSIox, phosphorus saturation index based on oxalate extraction • WTR, water treatment residuals
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Pierzynski and Gehl (2005) and Shober and Sims (2003) recommended using a P index rating used for confined animal feeding operations to determine if biosolids application should be based on N- or P-based agronomic rates (USDA, 2003). Many states have P index rating systems in place, including Colorado, which has developed a P risk index for confined animal feeding operations that accounts for differences in soil permeability, slope, soil-test P concentrations, P application rates, P application method, and best management practices (e.g., credits for reducing potential off-site P movement) (Sharkoff et al., 2005).
Best management strategies for land-applied biosolids could be based on total P content or P loading, but the biosolids-P form varies depending on the treatment process. Penn and Sims (2002) showed that biosolids produced using a biological nutrient removal process caused the greatest increase in extractable soil P and runoff-dissolved reactive P when added to soil. They further showed that biosolids produced with iron addition had the lowest P concentrations in these fractions. Biosolids soil amendment significantly increased Fe-P complexation, probably due to Fe addition during biosolids treatment, and a trend for greater Al-P complexation was evident where biosolids had been applied (Maguire et al., 2000). Maguire et al. (2001) showed that water extractable- and iron-strip extractable-P from soils amended with various types of biosolids followed the general trend of: soils amended with biosolids produced without the use of Fe or Al > soils amended with biosolids produced using Fe or Al and lime > soils amended with biosolids produced using only Fe and Al salts. The authors recommended testing biosolids for P availability, rather than total P, as a more appropriate indicator for predicting extractable P from biosolids-amended soils.
Predicting P bioavailability in biosolids-affected soils has typically been performed by characterizing fractions into labile and nonlabile pools. Sui et al. (1999) used a Hedley fractionation scheme on a biosolids-amended soil, noting a decrease in the proportion of HCl-P (a relatively resistant form) and a subsequent increase in NaHCO3–P (a relatively available form) concentration in the soil surface after biosolids amendment. The transformation was attributed to HCl-P dissolution as a result of decreased soil pH caused by biosolids addition. Maguire et al. (2000) used a six-step fractionation procedure on biosolids-amended soils and found a significant increase in content and percentage of Fe-bound P and a trend for greater Al-bound P. The authors further noted that desorbable P was closely related with Al-bound P, suggesting that this phase was the main source of desorbable P.
Our objectives were to assess the 20-yr impact of repeated, increasing biosolids applications on (i) the total recovery of P applied in biosolids, (ii) the dominant inorganic and relatively labile organic soil P phases, (iii) the degree of phosphorus saturation of surface soils, and (iv) the evaluation of the P index and risk assessment using the Colorado Phosphorus Risk Index.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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The soil was a Platner loam (fine, smectitic, mesic Aridic Paleustoll). Before biosolids application, soil organic matter content was <10 g kg–1 to a depth of 60 cm, surface (0–20 cm) soil pH was 6.5, and the pH of subsoil (20–60 cm) ranged from 6.6 to 7.5. Electrical conductivity of saturated-soil extracts were <1 dS m–1 at all soil depths, and NO3–N plus NH4–N was <7 mg kg–1 for the two depths (Utschig et al., 1986). The 4 M HNO3 extractable P concentrations (Bradford et al., 1975) in the 0- to 20-cm and the 20- to 60-cm depths were 0.48 and 0.62 g kg–1, respectively (Barbarick et al., 1986; Utschig et al., 1983).
Littleton/Englewood, Colorado, wastewater treatment facility (L/E) biosolids were generated by anaerobic digestion followed by approximately 2 mo of sand-bed drying. Biosolids samples were collected just before application and kept refrigerated at approximately 3°C until chemical analyses. After digestion with HClO4–HNO3–HF-HCl (Soltanpour et al., 1996), elemental composition of biosolids samples were determined using inductively coupled plasma–atomic emission spectroscopy (ICP–AES). Every 2 yr from 1982 to 2002, except in 1998, biosolids were applied at rates of 0, 6.7, 13, 27, or 40 Mg dry biosolids ha–1 to 3.6 by 17.1 m plots. Biosolids were not applied in 1998 because the land was intended to be sold to developers in that year. The plots treated with 40 Mg ha–1 received 6.7 Mg ha–1 in 1982 and 40 Mg ha–1 biosolids from 1984 through 1990. Beginning in 1992, we discontinued the 40 Mg ha–1 rate with the overall goal of determining the time required for agronomic parameters in these high-application plots to return to the control levels (e.g., received no biosolids or fertilizer application during the rest of the study). The effects of biosolids termination on agronomic parameters are published elsewhere (Barbarick and Ippolito, 2003). Four replications of all biosolids application rates in a randomized complete block arrangement were used. Biosolids were weighed (solids content ranging from 530 to 880 g kg–1) and corrected for moisture content, evenly spread over the plots using a front-end loader, hand raked to improve the uniformity of distribution, and incorporated to a depth of 10 to 15 cm with a rototiller.
Composite soil samples (three cores per plot) were collected from the 0- to 20-cm (tillage layer) and 20- to 60-cm depth near the center of each plot in July 2003. Samples were taken near the center of each plot to avoid potential biosolids redistribution problems that can occur after many tillage operations during many cropping years (Yingming and Corey, 1993). The soil samples were air-dried, crushed to pass a 2-mm sieve, and weighed for analyses. The research site was lost in 2004 to development.
Phosphorus Accountability
Yearly and cumulative masses of P applied for each application rate were calculated based on biosolids P content and biosolids load. Yearly grain samples were collected, digested with concentrated HNO3 (Huang and Schulte, 1985), and analyzed for P using ICP–AES. Yearly and cumulative masses of grain-P removed were determined based on P content and grain yield. Phosphorus contained within wheat straw was assumed to be returned to the soil during conventional tillage practices. The potential soil P accumulation was estimated as the difference between the amount of biosolids P added and the amount of P removed in grain. The actual increase in soil P (0- to 60-cm depth) was calculated from the difference between the background (i.e., 1982) and 2003 4 M HNO3 soil P concentrations.
Soil Phosphorus Fractionation
Inorganic P fractionation for noncalcareous soils was performed on 1.00 g of 0- to 20-cm depth soil according to methods outlined by Kuo (1996). The inorganic P extraction procedure identified (i) soluble/loosely bound P (1 M NH4Cl), (ii) Al-bound P (0.5 M NH4F), (iii) Fe-bound P (0.1 M NaOH), (iv) occluded P within the matrices of retaining components/minerals (0.3 M Na3C3H6O7 + 1 M NaHCO3 + Na2S2O4) (Evans and Syers, 1971), and (v) Ca-bound P (0.25 M H2SO4). Soils were washed and centrifuged twice with saturated NaCl between each step, with the NaCl solution added to the previous filtrate. Extracts were filtered through a 0.45-µm membrane before colorimetric P determination by spectrophotometer (882-nm wavelength), following a modified ascorbic acid procedure (Rodriguez et al., 1994). Modifications were made to the occluded P fraction due to insufficient color development. These modifications were based on previous research by Weaver (1974) whereby a 2.5-mL aliquot was transferred to a 25-mL volumetric flask. Then, 1.5 mL of 5% ammonium molybdate solution, 15 mL of deionized water, and 2.5 mL of color developing reagent (Rodriguez et al., 1994) were added, and the solution was brought to a final volume using deionized water. The final solution was allowed to stand for 30 min before P analysis.
Inorganic P fractionation for calcareous soils was performed on 1.00 g of 20- to 60-cm depth soil according to methods outlined by Kuo (1996). The inorganic P extraction procedure identified (i) soluble/loosely + Al-bound + Fe-bound P (0.1 M NaOH + 1 M NaCl), (ii) occluded P as previously outlined, and (iii) Ca-bound P (0.5 M HCl). Soil washing, centrifugation, colorimetric determination, and modifications to the occluded P fraction were followed as previously described.
Two organic P (Po) fractions were also identified: microbial biomass P (i.e., P originating from lysed microbial cells) and labile organic P (easily mineralizable Po) following procedures outlined by Zhang and Kovar (2000). Microbial biomass P was calculated as total labile P + CHCl3–lysed microbial cells minus total labile P. Labile Po was calculated as total labile P minus labile inorganic P. Briefly, total labile P was determined using 0.5 M NaHCO3 + K2S2O8, total labile P + CHCl3 was determined using 2 mL CHCl3 + 0.5 M NaHCO3 + K2S2O8, and labile inorganic P was determined using only 0.5 M NaHCO3. Phosphorus concentrations were determined colorimetrically as previously described.
Soil Phosphorus Saturation Index
Because the 0- to 20-cm depth was most influenced by biosolids application and incorporation, we analyzed soils from this depth and all treatments for oxalate extractable P, Al, and Fe (Loeppert and Inskeep, 1996) using ICP–AES. This soil depth was noncalcareous and had a pH of 6.6 (Ippolito and Barbarick, 2006). The oxalate-extractable P, Al, and Fe concentrations were used to determine the soil phosphorus saturation index (PSIox) (Schoumans, 2000):
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Based on findings from the 0- to 20-cm soil P fractionation experiment, we found Ca-P phases dominating in this system. We then determined the additional amount of P that could be sorbed by the 0- to 20-cm depth soil following a method outlined by Zhou and Li (2001). Three g soil + 30 mL of 50 mM KCl solution containing 400 µg P mL–1 (from K2HPO4) were placed into a 50-mL centrifuge tube, shaken for 24 h, and centrifuged, and the filtrate was passed through a 0.45-µm membrane (Zhou and Li, 2001). Phosphorus was determined on all solutions as previously described. The amount of P sorbed was determined by the difference in the initial and final-solution P concentrations.
Statistical Analysis
All data were tested for normal distribution, and, when appropriate, were log transformed (Steel and Torrie, 1980). All statistical tests were performed using the Proc GLM model in SAS software version 9.1 (SAS Institute, 2002). Differences within each P fraction were examined using ANOVA at 
= 0.05, with mean separation determined using Fisher's Protected LSD procedure.
Colorado Phosphorus Risk Assessment
Phosphorus risk assessment was conducted based on the Colorado Phosphorus Risk Assessment (Sharkoff et al., 2005) protocol, using the soil samples that received 6.7 and 13 Mg dry biosolids ha–1 application rates. These two rates of biosolids application represent the agronomic rate based on crop N requirement (Colorado Department of Public Health and Environment, 2003) and two times the agronomic rate, respectively. Soil-test P concentrations were determined using ammonium bicarbonate-diethylenetriaminepentaacetic acid (AB-DTPA) (Barbarick and Workman, 1987) because the AB-DTPA extraction test has historically been performed on soil from this research site.
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RESULTS AND DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Regressing the cumulative mass of P applied versus cumulative biosolids applied (up to the 27 Mg biosolids ha–1 application rate) gave the following equation:
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Using this equation and based on the predicted kg P displaced due to tillage (5.90 kg P), we found that approximately 41.4 kg biosolids ha–1, or 255 kg biosolids per plot, were redistributed over the plot lifetime. Essentially 4.14 kg biosolids ha–1, or 25.5 kg P per plot, per individual wheat-fallow rotation was transported across individual plots due to tillage. If the agronomic biosolids land application rate (6.7 Mg ha–1) were followed on a yearly basis, 0.95 kg P per plot, or 0.15 kg P ha–1, should be redistributed per year due to tillage. These values should aid future biosolids applicators in determining not only tillage P loss, but other nutrient and trace metal losses as well.
Soil Phosphorus Fractionation
Increasing biosolids application rate had no effect on the soluble/loosely bound (P = 0.129) or Ca-bound (P = 0.240) fractions in the 0- to 20-cm soil depth (Table 3). The soluble/loosely bound P concentration should be relatively low, compared with other fractions, regardless of application rate due to this fraction forming strong complexes with other soil phases present (i.e., Al-, Fe-, and Ca-bound phases). No difference in Ca-bound P was due to excess free Ca present in the system because this soil was derived from calcareous parent material. However, differences among treatments were evident for the Al-bound (P < 0.001), Fe-bound (P = 0.007), occluded (P = 0.002; log transformed), labile Po (P = 0.003), and biomass P (P = 0.003) fractions. The 27 Mg ha–1 rate contained the greatest P concentrations as compared with the other treatments. Maguire et al. (2000) used the same P fractionation technique and found that biosolids additions led to increases in Al- and Fe-bound P fractions when compared with untreated control soils.
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Biosolids used in our study averaged a total of 14.2 ± 5.4 g Fe kg–1, and with the addition of Fe2(SO4)3, an amorphous Fe phase was most likely present. Penn and Sims (2002) found that biosolids generated with Fe salt addition contained P mostly bound in the Fe-P phase within the biosolids. Nanzyo (1986) showed increased P adsorption on an amorphous iron hydroxide gel as compared with crystalline goethite, indicating that amorphous phases have a relatively large surface area for increased reaction to occur. As in our study, soils amended with biosolids containing amorphous Fe phases should contain increased Fe-bound P.
Increasing biosolids application rate had no effect on the soluble/loosely/Al-bound/Fe-bound (P = 0.453), Ca-bound (P = 0.616), or labile Po (P = 0.438) fractions in the 20- to 60-cm soil depth (Table 4). The majority of P in the 20- to 60-cm depth was bound to Ca, which was not surprising because the subsurface is calcareous and has an average pH of 8.1 (Ippolito and Barbarick, 2004). However, significant differences between treatments were evident for the occluded (P = 0.033) and biomass P (P = 0.038) fractions. The 13 and 27 Mg ha–1 rates contained the greatest accumulated occluded P compared with the untreated control. The increase in the occluded P phase suggests downward movement of Fe-bound P with a subsequent transformation to an occluded species or downward movement of the occluded phase. Our past research (Barbarick et al., 1997) used a 4 M HNO3 digest and found no significant biosolids-addition effects for any elements, including P, in the 20- to 60-cm soil depth at this site. In a follow-up study, Barbarick et al. (1998) found that AB-DTPA–extractable Zn increased with depth and biosolids application rate, illustrating the lack of sensitivity of 4 M HNO3 for discerning differences due to biosolids application rate. Thus, the utilized P fractionation scheme, by isolating each fraction separately, may be more sensitive to overall changes within each fraction.
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Barbarick and Ippolito (2003) noted that the decline in surface AB-DTPA–extractable P in the discontinued 40 Mg ha–1 rate could have been the result of residual organic matter mineralization leading to plant P uptake, formation of slightly soluble soil minerals or adsorption onto soil minerals, or leaching into the subsoil, thus reducing their lability in surface soil. Grain P removal was relatively negligible (1.5% of the P applied on the 27 Mg ha–1 plots; Table 1), and thus the sequential extraction results lend support that P is forming slightly soluble mineral phases, is sorbed to mineral surfaces, or is leached. Soil P in the 0- to 20-cm depth formed strong associations with Fe phases as noted in the Fe-bound and occluded phases and to some extent in the Al-bound phases (Table 3). Although Barbarick et al. (1997) showed significant quadratic or exponential rise in AB-DTPA–extractable P contents with increasing biosolids applications, the AB-DTPA extractant is used to identify Ca-bound P phases. Therefore, given the time since discontinuance, the 40 Mg ha–1 application P content most likely transitioned from a Ca-bound to a Fe/Al-bound form (Table 3).
In addition to being bound to less soluble AB-DTPA–extractable forms in the 0- to 20-cm depth, occluded P concentration increased with increasing biosolids application rate in the 20- to 60-cm depth (Table 4). This P was likely leached from the soil surface over the 20-yr study period, as suggested previously by Barbarick and Ippolito (2003).
Of the two organic fractions studied, labile Po dominated the soil surface, biomass P dominated the subsurface, and both fractions increased with increasing biosolids rate (Tables 3 and 4). Biomass P decreased with increasing biosolids application in the 0- to 20-cm depth. A combination of severe drought from 1999 to 2003, conventional tillage, biyearly and increasing biosolids application rates, crop failure in 2000 and 2002, and soil sampling timing may have resulted in microorganism dormancy or mortality in the soil surface. Srivastava (1998) showed that air-drying soils decreased microbial biomass C but increased extractable P, to some extent due to microbial biomass death due to air drying. Tate et al. (1991) observed constant microbial biomass, but microbial P content varied seasonally, with Po mineralization in spring and immobilization in the early winter months. Biomass reduction in the soil surface, via mortality or dormancy, may have been further exacerbated by drought conditions. Lynch and Panting (1980) found that soil microbial biomass increased during the wheat-growing season and then decreased to an approximately constant amount. Increased microbial activity was presumably related to increased soil temperatures and crop growth, with maximum activity in June during grain heading and decreasing to a constant amount by August. Maximum biomass coincided with maximum root production (Lynch, 2002). Poor crop growth in our system over several years, in combination with the aforementioned negative variables, could have reduced surface microbial activity. Furthermore, the sown crop was most likely stressed, and when seed formation began it required the mobilization of P to seed grain, thus placing more demand on available soil P. Microbial biomass P should rapidly mineralize and release P to satisfy seed formation requirements (Agbenin and Adeniyi, 2005), yet surface microbial activity was lessened, further stressing the crop.
Subsurface microbial biomass was greater than the surface and increased with increasing biosolids application rate (Table 4). This was possibly due to protection from environmental and anthropogenic variables, such as drought and conventional tillage practices, or was increased due to increases in dissolved organic carbon. Decreased anthropogenic activities (e.g., reduction in tillage practices) increase microbial biomass more in permanent than annual pastures and more in permanent grass than arable soil (Brookes et al., 1984; Milne and Haynes, 2004). Chen et al. (2004) attributed increased microbial activity to greater concentrations of water-soluble soil organic carbon. Using soil from the 0- to 20-cm depth of the 27 Mg ha–1 rate from our study site, Al-Wabel et al. (2002) found increased dissolved organic carbon in effluent from undisturbed soil columns as compared with control (no biosolids application) soil. These findings help explain the observed increase in subsurface microbial biomass in our system.
Soil Phosphorus Saturation Index
Because biosolids contain inherently large quantities of P, there is a risk of P saturation in treated soils. The potential risk of off-site P movement can be determined using oxalate-extractable Al, Fe, and P concentrations. The Al, Fe, and P oxalate-extractable concentrations increased with biosolids addition up to the 27 Mg ha–1 application rate (Fig. 1A–C). The 40 Mg ha–1 application rate, which received a cumulative biosolids application similar to the 13 Mg ha–1 rate, had Al, Fe, and P concentrations similar to the 6.7 and 13 Mg ha–1 rates. This is comparable to the P fractionation findings and those of Barbarick and Ippolito (2003). Maguire et al. (2000) found that Fe and P oxalate-extractable concentrations were greater in biosolids-amended as compared with control soils.
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Based on findings from the 0- to 20-cm soil P fractionation experiment, we noted that Ca-P phases, in addition to Fe-P phases, were dominating this system. Thus, we determined the soil P sorption characteristics using a method outlined by Zhou and Li (2001). No significant difference (P > 0.05) existed between biosolids treatments, and overall the soil had the ability to sorb from 760 to 1040 mg kg–1 of additional P (data not shown). In support of this finding, the average (1982–2002) biosolids P/Fe molar ratio was 2.9, suggesting that the biosolids did not contain enough Fe to bind P completely, and therefore the formation of other soil P phases could be expected.
Increasing the soil P sorbing capacity by the addition of P-sorbing materials should be considered for fixing excess soil P, preventing off-site movement and subsequent waterway eutrophication. Ippolito and Barbarick (2006) added increasing amounts of Al-based water treatment residuals (WTR) to the 6.7 Mg biosolids ha–1 treated soil described in this paper and showed that, at a ratio of 4:10 WTR/soil or greater, water-extractable P significantly decreased to near detection limits (0.04 mg P L–1). Novak and Watts (2004) mixed WTR with high P–bearing soil from long-term manure applications and noted a several-fold increase in the P sorption maximum as compared with soil alone. Dayton and Basta (2005) added Al-based WTR to a soil with a history of poultry litter application and found a significant reduction in 0.01 M CaCl2 extractable P content. They suggested promoting WTR use as a best management practice to reduce P loss from agricultural land. Thus, addition of P-sorbing materials to systems similar to that discussed in our research can prolong their life expectancy regarding long-term biosolids application and P-loading.
Colorado Phosphorus Index Risk Assessment
To help manage manure application and to comply with the Natural Resources Conservation Service Nutrient Management Conservation Practice Standard, Code 590, Sharkoff et al. (2005) developed the Colorado Phosphorus Index risk assessment. Although not included in this risk assessment, Shober and Sims (2003) and Pierzynski and Gehl (2005) have recommended such a technique for managing biosolids relative to P applications. Instead of using one extraction technique, such as water-extractable P to identify potential offsite movement, the current consensus among states is to follow a comprehensive systems approach in identifying P loss from agricultural fields (Elliott et al., 2005). Colorado uses an additive approach; Table 5 provides the overall Colorado P Index Risk assessment based on the total score of the aforementioned factors. Table 6 provides the estimated risk index for the research site assuming a recommended agronomic rate of biosolids is applied (i.e., 6.7 Mg ha–1), which is the likely scenario for biosolids land application. Table 6 also includes results for excessive biosolids application (13 Mg ha–1) for comparison.
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If biosolids are to be included in future risk indices, a complete biosolids analysis should be considered, especially regarding Fe and Al content. Elliott et al. (2005) stated that the current Pennsylvania P index groups biosolids and manures into a single category, yet their research showed that most biosolids studied had runoff P losses less than dairy manure. In the case of Philadelphia, Pennsylvania biosolids, Fe content was exceptionally high (72 g kg–1) due to disposal of Fe-based WTR into the sanitary sewer system. Forty-seven states have adopted a P index approach, yet only nine states use estimates or weighting factors for determination of Po-applied availability (Sharpley et al., 2003). In terms of best management practices, Elliott et al. (2005) suggested using P source coefficients for individual biosolids because the susceptibility of P to leaching and runoff is variable, and without source coefficients site vulnerability risk could be compromised. Biosolids source coefficients would reflect the portion of total applied P susceptible for off-site transport (Elliott et al., 2005). Other best management practices could include the addition of P-sorbing materials (e.g., WTR) at time of application.
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CONCLUSIONS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Of the two organic fractions studied, labile Po dominated the soil surface, whereas biomass Po dominated in the subsurface. Biomass P decreased and increased with increasing biosolids application in the 0- to 20 and 20- to 60-cm depths, respectively. This suggests that (i) a combination of severe drought from 1999 to 2003, conventional tillage, repeated biosolids application, crop failure in 2000 and 2002, and soil sampling timing may have forced microorganism dormancy or mortality in the soil surface; (ii) biomass P reduction in the soil surface, regardless of the process, may have been further exacerbated by drought conditions; and (iii) subsurface microbial biomass was greater than surface microbial biomass, possibly due to protection from environmental and anthropogenic variables, such as drought and conventional tillage practices, or was increased due to an increase in dissolved organic C influx.
Long-term, repeated biosolids application caused increased PSIox values in the soil surface (0- to 20-cm depth) as compared with the control regardless of application rate, with values ranging from 60 to almost 100%. However, the P fractionation scheme showed that Ca-P phases were also present in large quantities, suggesting that further P sorption could occur. Determination of the soil surface P sorptive capacity showed that this system can still sorb an additional 760 to 1040 mg P kg–1, suggesting that additional soil P precipitation could occur. The average biosolids P/Fe molar ratio indicated that biosolids did not contain enough Fe to bind P completely, and therefore the formation of other soil P phases would be expected.
From a risk-management perspective, continuous biosolids application can be based on the required crop N agronomic rate. Biosolids over-application, as shown by applying twice the agronomic rate every other year over 20 yr, increased the risk for off-site P movement, and thus biosolids application would be limited based on crop P requirements, not N. Once regulatory limits are reached at a particular location, operators would need to cease application for about 6 yr to allow soil AB-DPTA P levels to drop significantly below the limitations. Proper monitoring and management would ensure that off-site P movement does not become a major issue.
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