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Published in J. Environ. Qual. 32:1773-1781 (2003).
© 2003 ASA, CSSA, SSSA
677 S. Segoe Rd., Madison, WI 53711 USA

TECHNICAL REPORTS

Surface Water Quality

Biosolids Decomposition after Surface Applications in West Texas

W. F. Jaynes*,a, R. E. Zartmana, R. E. Sosebeeb and D. B. Westerb

a Plant and Soil Science Department, Texas Tech University, Lubbock, TX 79409
b Range, Wildlife, and Fisheries Management Department, Texas Tech University, Lubbock, TX 79409

* Corresponding author (william.jaynes{at}ttu.edu).

Received for publication June 2, 2002.

   ABSTRACT
TOP
 NOTES
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
In a semiarid environment, climate is a critical factor in the decomposition of surface-applied biosolids. This study examined the effect of 2- to 7-yr exposure times on the composition of single applications of New York, NY biosolids in western Texas. Exposure time effects on organic matter, N, P, S, Cu, Cr, Pb, Hg, and Zn were studied near Sierra Blanca, TX. Due to organic matter decomposition, total organic C decreased from 340 g kg-1 in fresh biosolids to 180 g kg-1 in biosolids after 82 mo of exposure, whereas the inorganic ash content of the biosolids increased from 339 to 600 g kg-1. Total N decreased from 50 to 10 g N kg-1 and total S decreased from 12 to 6 g S kg-1. Bicarbonate-available P in the biosolids decreased from 0.9 to 0.2 g kg-1. Successive H2O extractions yielded soluble P concentrations consistent with dicalcium phosphate (dical) for fresh biosolids and tricalcium phosphate (trical) for biosolids exposed for 59 months or more. Sparingly soluble phosphates, such as dical and trical, potentially yield >0.5 mg P L-1 in runoff waters for extended periods after biosolids applications, especially after multiple applications. Selective dissolution of the biosolids indicated that as much as 66 to 78% of P exists as iron phosphates, 16 to 21% as Fe oxides, and 5 to 12% as insoluble Ca phosphates. Chemical analyses of ash samples suggest that Cu and Zn have been lost from biosolids through leaching or runoff and no losses of Pb, Cr, or Hg have occurred since application.

Abbreviations: dical, dicalcium phosphate • monocal, monocalcium phosphate • octocal, octocalcium phosphate • PSRP, process to significantly reduce pathogens • trical, tricalcium phosphate


   INTRODUCTION
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
MANY STUDIES have examined the effects of the application of biosolids (anaerobically digested sewage sludge) on soil properties, plant productivity, and runoff water quality (Ryan et al., 1973; Terry et al., 1979; Benton and Wester, 1998; Stehouwer et al., 2000; Yan et al., 2000; Rostagno and Sosebee, 2001). Biosolids are typically mixed into the surface layers of cultivated soils through tillage operations. Direct analyses of the changes in biosolids composition after incorporation are not practical. Post-application changes in biosolids composition, therefore, must be inferred from changes in soil composition.

Sommers et al. (1976) concluded that, although there was a seasonal variation, the organic carbon content of sewage sludge was relatively constant with respect to sampling time. They also reported that ash contents of biosolids from medium-sized cities (populations of 5000–76000) ranged from 46 to 57%. Ash is the inorganic residue left after the organic matter in a sample is burned off in a furnace at temperatures of >500°C. Tester et al. (1977) measured a 16% loss of C as CO2 in sewage sludge compost after incubation for 54 d. Terry et al. (1979) reported that organic matter decomposition was greatest in surface-applied biosolids. Sosebee et al. (1993) reported that organic matter content decreased from 55 to 42% within six months after topical biosolids application.

Sommers (1977) determined that the median N concentration for anaerobically digested sewage sludge was 42 g kg-1 and that most of total N was organic (Sommers et al., 1976). Parker and Sommers (1983) measured a 15% mineralization of N in biosolids during a 16-wk incubation. Similarly, Rostagno and Sosebee (2001) reported that total N decreased 35% one year after topical biosolids application.

Sposito et al. (1982) noted that the humic acids separated from biosolids have sulfur-containing anionic surfactants derived from household detergents. Holt et al. (1989) examined the fate of linear alkylbenzene sulfonates (i.e., from detergents) in biosolids-amended soils and reported losses of >98% at most sites, which they attributed to microbial breakdown rather than leaching. Boyd and Sommers (1990) similarly reported that sodium lauryl sulfate and other anionic surfactants impart a higher S concentration to fulvic acid fractions in biosolids.

Boyd and Sommers (1990) measured humic and fulvic acid contents in some biosolids and in biosolids-amended soils. They noted that humic and fulvic acid fractions from biosolids contained higher N contents, lower C to N ratios, higher H to C ratios, and lower carboxyl group acidity than soil-derived humics. The biosolids humic fractions were reported to contain 30 to 50% nonhumic substances including amino acids, hexosamines, neutral sugars, and anionic surfactants.

Soon and Bates (1982) reported that about 70% of the inorganic P in biosolids treated with Al2(SO4)3 or FeCl3 was extracted by NaOH. Aluminum and ferric iron salts are commonly added in fresh water and wastewater treatment to precipitate P and sediment suspended solids. Robertson (2000) reported that Fe from iron oxides or oxyhydroxides is reduced during anaerobic sludge digestion and it reacts with dissolved P to form insoluble Fe phosphates. Another Fe source would be needed to form Fe phosphate because human feces and urine do not contain much Fe. Human adults excrete no more than 6.8 mg Fe d-1, whereas as much as 2770 mg P d-1 is excreted (Lentner, 1981).

Maguire et al. (2000) reported that biosolids applied in Maryland, Delaware, and Virginia increased oxalate-extractable Fe and Al in the soils. They concluded that the oxalate-extractable Fe and Al might act to mitigate P release into runoff. Most of the biosolids studied by Maguire et al. (2000), however, had been treated with Fe salts during the wastewater treatment process. As noted by Robertson (2000), added Fe salts would precipitate as Fe phosphates during wastewater treatment. The P in Fe phosphates, such as meta-strengite, would not be as available to plants as dicalcium phosphate (dical), a more soluble phosphate form.

Olsen and Sommers (1982) noted that P fractionation methods yield uncertain results for heavily fertilized soils because intermediate reaction products (i.e., dical, octocalcium phosphate [octocal], tricalcium phosphate [trical]) with unknown dissolution characteristics can persist for several years. These phosphates were termed intermediate based on research by Lindsay (1979) and others on the dissolution and precipitation reactions that converted fertilizer phosphorus (i.e., monocalcium phosphate [monocal]) to more stable phosphorus minerals in soils. These phosphates had a solubility that was intermediate between the very soluble monocal and the insoluble apatites.

Sharpley and Moyer (2000) concluded that water-extractable P might be used to estimate the potential of land-applied manure or compost to contribute to P in runoff. They reported a strong correlation between P leached in five simulated rainfall events to P dissolved in water using a modified Hedley et al. (1982) fractionation. The modification of the Hedley et al. (1982) fractionation used by Sharpley and Moyer (2000) used a single 200-mL H2O extraction of a 1-g soil sample to measure soluble P. A single H2O extraction might effectively remove very soluble phosphates, such as monocal. No existing selective dissolution or extraction method, however, could adequately measure all of the dissolved P that sparingly soluble phosphates, like dical, octocal, and trical, would probably contribute to runoff. If a solid phase, such as dical, controls P solubility, soluble P concentrations will be essentially constant through many H2O extractions or rainfall events until completely dissolved.

MERCO Joint Venture LLC, a State of New York limited liability company, operated a biosolids management and beneficial land application program in Hudspeth County, Texas near the city of Sierra Blanca (MERCO, 2000). From 1992 to 2001, biosolids from New York City municipal sewage treatment plants were applied to the surface of native range soils at the site. Most of the biosolids received at the MERCO site from 1992 to 1998 were treated in New York to reduce pathogens by a method termed Process to Significantly Reduce Pathogens (PSRP). From 1998 to 2001, most of the biosolids received at the MERCO site were anaerobically digested for a shorter time interval than the PSRP material and had to be treated with quicklime (CaO) at the site to meet the pathogen reduction standards for alkaline-stabilized sewage sludge (MERCO, 2000).

New York City operates 14 wastewater treatment plants that together generate about 1.2 Gg (0.3 dry Gg) of dewatered sewage sludge per day. From 1938 until 1992, liquid sewage sludge was taken by barge from the treatment plants and dumped in the Atlantic Ocean. The federal Ocean Dumping Ban Act of 1988 ended this method of sludge disposal (Department of Environmental Protection, 1998a). In 1992, New York City ended ocean dumping and started a sludge (biosolids) recycling program. The sludge was anaerobically digested at 35°C for 15 to 20 d (PSRP process) and subsequently dewatered to produce biosolids. Approximately 6% of the biosolids failed to meet quality control requirements for beneficial use and were deposited in landfills (Department of Environmental Protection, 1998b). Of the 94% of the biosolids that could be recycled, 39% was dried into fertilizer pellets, 12% was composted and used as a mulching material, and 8% was alkaline-stabilized and used as an agricultural liming agent. The remaining 35% of the biosolids was used for direct land applications (Department of Environmental Protection, 1999). The biosolids from NYC were largely residential in origin; the industrial portion of plant metal influent was reduced from approximately 30% in 1973 to approximately 5% in 1992 and remained near 5% through 1998 (Department of Environmental Protection, 1998c).

A number of research studies at the MERCO site have examined the effects of surface applications of biosolids on plant productivity, soil water and air quality, and soil chemical and physical properties (Harmel et al., 1997; Benton and Wester, 1998; Strait et al., 1999; Rostagno and Sosebee, 2001). A commercial biosolids application rate of 7 Mg ha-1 (dry wt.) was initially approved by the Texas Natural Resources Conservation Commission (TNRCC). The TNRCC later increased the rate to 18 Mg ha-1, which could be applied up to two times per year by MERCO. Earlier studies at the site determined that biosolids applications to the soil surface reduced soil erosion and minimized the transport of biosolids components. In contrast, mixing biosolids into the soil by tillage operations increased soil erosion and enhanced the movement of biosolids components in runoff (Fish and Shanks, 1997). Tillage has a long-lasting effect on vegetative cover and erodibility in an arid climate. However, Kleinman et al. (2002) studied the effects of fertilizer and manure P sources on runoff water P concentrations and found that significantly more P was lost from surface-applied P sources than from P sources mixed into the soil.

The MERCO site has a semiarid climate with 310 mm yr-1 precipitation, a mean annual temperature of 18°C, and an elevation of 1350 m above mean sea level (Rostagno and Sosebee, 2001). Most soils at the MERCO site are in an aridic soil moisture and a thermic soil temperature regime. In this climate, the biosolids materials are very persistent and large aggregates can readily be identified and sampled many years after application. For this study, PSRP biosolids samples with a range in exposure ages were collected at the MERCO site. The objectives of this research were to (i) measure the chemical and mineralogical compositions of the biosolids and evaluate the changes in composition that have occurred during exposure on the soil surface and (ii) relate the changes in composition to the mobility and probable fate of nutrient elements (such as N, S, and P) and potentially toxic elements (such as Cu, Zn, Cr, Cd, As, Hg, and Pb).


   MATERIALS AND METHODS
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Samples of fresh biosolids and biosolids applied 2 to 7 yr before sampling were collected at the MERCO site. Fresh biosolids were collected from a temporary holding shed at the MERCO site and 2- to 7-yr-old samples were collected from the soil surface at various locations on the site. The sampling sites had received only one application of PSRP biosolids. At the 1993 site, biosolids were applied to open rangeland, whereas biosolids at the 1992, 1994, and 1997 sites were applied to enclosed 0.5- to 7-m2 research plots. The equipment used to distribute biosolids on the open rangeland site yielded many large (>100 mm in diameter) readily identifiable aggregates. Biosolids at the research plots were hand-applied. The larger biosolid aggregates were sampled at both the rangeland and research plot sites to minimize contamination by local soil materials. Extraneous soil and other materials were removed from the biosolids sample aggregates (approximately 5–50 mm in diameter) during field collection. The biosolids aggregates were later agitated on a 2-mm sieve to remove any remaining extraneous material. Biosolids aggregates were subsequently ground to <2 mm using a coffee grinder.

Carbon Components
Total organic carbon was determined by wet digestion with chromic acid using a modified Mebius (1960) procedure (Nelson and Sommers, 1982; Method 29-3.5.3). In this method, biosolids samples were boiled with 0.067 M (0.4 N) chromic acid and the remaining unreacted chromic acid was measured by back-titration with ferrous ammonium sulfate.

Humic acid was fractionated from whole biosolids samples by extracting with 0.5 M NaOH followed by precipitation after adjustment to pH 1 or 2 by the addition of 12 M HCl. The precipitated humic acid was washed by centrifugation, dried, and weighed (Schnitzer, 1982). The fulvic acid fraction was determined by measuring the organic carbon contents in the supernatants after precipitation of humic acid. The supernatants contained organic matter soluble in both alkaline and acid solutions. Boyd and Sommers (1990) termed this alkaline- and acid-soluble fraction as the fulvic acid fraction. Separation of pure fulvic acid from lower molecular weight organics requires techniques such as adsorption to and elution from cation exchange resins (Schnitzer, 1982).

Total polysaccharides were measured in the biosolids using the method of Lowe (1993). In the total polysaccharide method, all polysaccharides in a sample were hydrolyzed to saccharide monomers by treatment with 12 M H2SO4 followed by autoclaving. Total polysaccharides include high molecular weight materials such as cellulose. A yellow-colored complex formed using phenol–sulfuric acid reagent was prepared to estimate total sugar content in the hydrolyzed samples. Glucose was used as a standard to calculate the sugar content based on the absorbance of the colored complex at 490 nm.

Nitrogen and Sulfur Analysis
Total N was extracted from whole biosolids samples with boiling concentrated H2SO4 (i.e., total Kjeldahl N digestion; Bremner and Mulvaney, 1982) followed by total N determination on the extracts using a Hach Company procedure and chemicals (Method 10071; Hach Company, 1996). The method used an alkaline persulfate digestion to convert all forms of N to nitrate. The nitrate was subsequently reacted with chromotropic acid to form a yellow complex that was measured using a colorimeter at 410 nm.

Total S in whole biosolids was oxidized to SO4 by dry-ashing 2 g of biosolids mixed with 1 g of NaHCO3 in a porcelain crucible at 550°C (Tabatabai, 1982). The ash was dissolved in 1 M HCl and excess BaCl2 was added to precipitate SO4 as BaSO4. The precipitated BaSO4 was filtered, dried, and weighed to calculate total S.

Reference Phosphates
Standard phosphate minerals were collected for use as references in P extractions of biosolids. Natural samples of variscite (AlPO4·2H2O) and fluorapatite [Ca5(PO4)3F] were obtained from Minerals Unlimited (Ridgecrest, CA). Synthetic ß-tricalcium phosphate [Ca3(PO4)2] was prepared by calcining a mixture of Ca2(HPO4)2 and Ca(OH)2 using a 1:1 molar ratio at 1050°C (Jaynes et al., 1999). Meta-strengite (FePO4·2H2O), dicalcium phosphate [Ca2(HPO4)2], and monocalcium phosphate [Ca(H2PO4)2·H2O] were obtained from Fisher Scientific (Pittsburgh, PA).

Phosphorus Analysis
Total orthophosphate P was determined using 0.5 M H2SO4 extraction of whole biosolids samples that had previously been heated at 550°C. Similarly, inorganic orthophosphate P was determined by extraction of unheated whole biosolids samples with 0.5 M H2SO4 (Olsen and Sommers, 1982; Method 24-3.3). By this method, the difference between total and inorganic orthophosphate is attributed to organic P. The extracts were saved for subsequent P analysis.

Available P was measured by extraction with pH 8.5, 0.5 M NaHCO3 (Olsen and Sommers, 1982; Method 24-5.4). Duplicate 1.0-g samples of biosolids and 0.1-g samples of reference phosphates were weighed into 50-mL centrifuge tubes and 20 mL of pH 8.5, 0.5 M NaHCO3 were added. The 0.1-g samples of the reference phosphates contained about the same total P contents as the 1.0-g biosolids samples. The tubes were agitated on a reciprocating shaker for 30 min and centrifuged, and the supernatants were saved for later P analysis.

Successive Phosphorus Extractions
Biosolids samples (1.0 g) and reference Ca-phosphates (0.1 g) were successively extracted five times with 50-mL aliquots of distilled water to compare the P solubilities and identify the most likely forms of Ca phosphate present in the biosolids. Duplicate biosolids and reference phosphate samples were weighed into centrifuge tubes and agitated on a shaker for 22 to 24 h to ensure dissolution of soluble P. The tubes were centrifuged and the supernatants retained in vials. The process was repeated four more times until a total of five extracts was collected.

Sequential Phosphorus Extraction
Sequential P extractions were performed on 1-g whole biosolids samples and 0.1-g reference samples of Ca-, Al-, and Fe-phosphates. The extractions were 0.1 M NaOH and 1.0 M NaCl (NaOH), sodium citrate–sodium bicarbonate (CB), citrate–bicarbonate–dithionite (CBD), and 1 M HCl (Olsen and Sommers, 1982; Method 24-4.2). By this method, the P in the NaOH + CB extractions was attributed to nonoccluded P, such as that in Fe- and Al-phosphates. The P in the CBD extract was attributed to P occluded in Fe oxides and oxyhydroxides. The HCl-extractable P was attributed to insoluble Ca phosphates, such as hydroxyapatite and fluoroapatite. The molar ratios of Fe to P and Ca to P were determined using a different sequential extraction that used a 0.5 M NaOH extraction followed by extraction with 1 M HCl; the ratios were based on the concentration of P in both extracts and the Fe and Ca concentrations in the HCl extract. Iron and Ca were precipitated in the NaOH extraction. Iron was determined by colorimetry and Ca was measured by atomic absorption spectrometry. Orthophosphate P was determined by colorimetry on all solutions using a modified Murphy and Riley (1962) method (Olsen and Sommers, 1982; Method 24-3.4).

Inorganic Analysis
Whole biosolids samples were weighed in duplicate into porcelain crucibles and heated to approximately 700°C for 1 h to remove organic materials. The resulting ash samples were weighed to determine biosolids ash contents. Duplicate samples (100 mg) of biosolids ash were weighed into teflon-lined, steel decomposition bombs and 6 mL of HF and 1 mL of aqua regia were added. After 1 h heating at 110°C, the sealed bombs were opened and the digests were transferred to 100 mL of a 56 g L-1 boric acid solution (Bernas, 1968). After complete dissolution, the solutions were diluted to 250 mL and transferred to plastic bottles. The solutions were later analyzed for total Si, Al, Fe, Ca, Mg, Na, K, P, Mn, Zn, Cu, Cd, As, Cr, Hg, and Pb using inductively coupled plasma spectroscopy.

Mineralogy
Whole biosolids powder samples were packed into aluminum sample holders and scanned from 2 to 40°2{theta} using CuK{alpha} ({lambda} = 1.5418 Å) radiation and a Phillips (Eindhoven, The Netherlands) X-ray diffractometer. Dominant mineral phases in the biosolids were identified by comparing peak positions and relative intensities in the diffraction patterns to reference values in the powder diffraction file (Joint Committee on Powder Diffraction, 2001).


   RESULTS AND DISCUSSION
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
CONCLUSIONS
 REFERENCES
 
Initial chemical analyses were obtained for biosolids samples received by MERCO during the month that biosolids were applied at the study sites (Table 1). The micronutrient element concentrations decreased yearly with mean values for 1992 through 1994 that exceed the values of 1997. These decreases mirror the reduced total metal (Ag, As, Cd, Cr, Cu, Hg, Pb, Ni, and Zn) loading reported by the Department of Environmental Protection from >74100 kg d-1 in 1973 to <1800 kg d-1 in 1992 and further reduction to approximately 1400 kg d-1 in 1998 (Department of Environmental Protection, 1998c). Similarly, Stehouwer et al. (2000) examined the chemistry of sewage sludge applications in Pennsylvania over a 20-yr period (i.e., 1978–1997) and reported large decreases in Cd, Cr, Cu, Pb, Hg, Ni, and Zn.


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  		Table 1. Representative initial compositions (average and standard deviation) of biosolids.

 
Carbon Components
Total organic carbon (TOC) values decreased from 340 g kg-1 in fresh biosolids to 180 g kg-1 in biosolids exposed for 82 mo (Fig. 1a) . Ash contents increased from 339 to 600 g kg-1 corresponding to a great loss of organic matter and a residual increase in inorganic material. Sosebee et al. (1993) reported that organic matter content decreased from 55 to 42% within 6 mo after biosolids application at the MERCO site. Similarly, Terry et al. (1979) noted high organic matter decomposition in surface-applied biosolids.



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  		Fig. 1. (a) Contents of ash, total organic carbon (TOC), (b) total polysaccharides, total nitrogen, and total sulfur in biosolids. Vertical error bars indicated on points where errors exceed limits of point marker.

 
Humic acid decreased from 133 to 67 g kg-1 in the 0- to 82-mo-old biosolids, whereas the fulvic acid fraction decreased from 164 to 112 g kg-1 C. Humic acid and fulvic acids are soluble forms of organic matter that form soluble complexes with heavy metals. Holtzclaw et al. (1977) observed that almost all of the Cu in biosolids was associated with humic acid. Total polysaccharide contents (Fig. 1b) were slightly less in the biosolids with a greater exposure age. Decreases in total organic C were greater than were decreases in polysaccharides. Hence, the relative content of polysaccharides in the organic matter increased with exposure age. The greater proportion of polysaccharides in the residual organic matter explains the fibrous appearance of the older biosolids samples.

Nitrogen and Sulfur Analysis
Total nitrogen decreased from 50 g kg-1 in fresh biosolids to 10 g kg-1 in the 82-mo biosolids, whereas total sulfur decreased from 12 to 6 g kg-1 (Fig. 1b). Zartman (1993) noted that total Kjeldahl N concentrations were only 0.7 g kg-1 in soils at the MERCO site before biosolids applications. Hence, biosolids can significantly increase soil N contents. A fraction of the N decreases with exposure were due to ammonia volatilization. Harmel et al. (1997) modeled ammonia volatilization at the MERCO site and measured volatilization losses that ranged from 11.5 kg NH3–N ha-1 in the cool-season trial to 35.5 kg NH3–N ha-1 in the hot-season trial. Tabatabai and Chae (1991) leached soils amended with sewage sludge for 26 weeks and measured a 44 to 97% sulfur mineralization. Apparently, the semiarid environment at the MERCO site can only support a much lower sulfur mineralization rate. Decreases in N and S paralleled decreases in organic carbon and probably reflect the decomposition of organic matter. Organic polymers such as polyacrylamide were added at the sewage treatment dewatering plants (Department of Environmental Protection, 1998b). These materials probably also contributed to the N, S, and C contents of the biosolids.

Phosphorus Analysis
Total and inorganic P contents of the biosolids showed little change, whereas available P contents decreased with exposure (Table 2). The available fraction of inorganic P in biosolids relative to the reference phosphates suggests that the P in fresh biosolids (0.045 = 0.9/20.1) is comparable in availability to dical (0.037 = 8.5/229.5). Olsen and Sommers (1982) reported that available P in soils measured to be >0.01 g kg-1 by their method have sufficient P for plants and a yield response would not be expected from P fertilizer applications. Available P concentrations even in the biosolids samples with the greatest exposure age were 20 times greater than 0.01 g kg-1. However, this does not reflect available P in the whole soil. Biosolids mixed with a hectare-furrow-slice of soil at the 90 and 18 Mg ha-1 application rates would dilute available P by 1:25 and 1:125, respectively, and available P concentrations in the whole soil would be at or below the sufficiency level.


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  		Table 2. Total P, inorganic P, available P, and water-soluble P concentrations in samples.{dagger}

 
Successive Phosphorus Extractions
The results of the successive water extractions indicated a decrease in P solubility with exposure (Table 2). In the first and fifth extractions of the fresh (0 mo) biosolids sample, dissolved P values decreased from 26 to 6 mg L-1. In contrast, soluble P in the oldest biosolids sample only decreased from 2 to 1 mg L-1. Soluble P values for the reference dical and trical samples were much less variable with dical at approximately 6 mg L-1 and trical at approximately 1 mg L-1. The soluble P values in the last extraction should best reflect the control of soluble P by phosphates in the biosolids. The approximate 6 mg P L-1 for the fresh biosolids suggested the presence of a solid phase, such as dical. The approximate 1 mg P L-1 in the 82-mo-old biosolids suggested that trical might control P solubility. Whatever solid phases actually controlled the P solubility, about 18% (i.e., 3.7/20.1) of total inorganic P was soluble in the fresh biosolids, whereas only approximately 2% (i.e., 0.4/22.4) was soluble in the biosolids after 59 mo or more exposure.

The five 50-mL H2O extractions (Table 2) dissolved <7% of the P in the dical samples. Based on these extractions, a 0.1-g dical sample could yield a dissolved P concentration of approximately 6 mg L-1 in more than 3.7 L of H2O. Similarly, a 0.1-g sample of trical could yield a dissolved P concentration of approximately 1 mg L-1 in more than 18.7 L of H2O. Manahan (2000) noted that algae grew at PO3-4 concentrations as low as 0.05 mg L-1 (i.e., P = 0.016 mg L-1). Hence, the presence of sparingly soluble phosphates, such as dical and trical, in soils or biosolids can potentially maintain dissolved P concentrations of >0.02 mg L-1 in runoff waters.

Simple calculations can be used to illustrate how the presence of sparingly soluble phosphates can support relatively high P concentrations in runoff. A 25-mm rainfall event is equal to 2.5 x 108 cm3 H2O ha-1, which is equal to 13.9 cm3 H2O g-1 of biosolids for an 18 Mg ha-1 application, but only 2.8 cm3 H2O g-1 of biosolids for a 90 Mg ha-1 application. Each of the five extractions in Table 2 used 50 mL of H2O g-1 of biosolids. An annual precipitation of 310 mm yr-1 for 18 mo is, at most, equivalent to five 50-mL water extractions at the 18 Mg ha-1 rate and just one 50-mL water extraction at the 90 Mg ha-1 rate. However, a shorter contact time and poorer mixing would make rainfall less effective in dissolving P in biosolids than the water extractions.

At the MERCO site, Rostagno and Sosebee (2001) measured a PO4–P concentration in runoff from simulated rainfall (80 mm in 30 min) of 4.96 mg L-1 two weeks after a biosolids application of 90 Mg ha-1 to the Stellar soil (fine, mixed, superactive, thermic Ustic Calciargid). The PO4–P concentration, however, only decreased to 4.31 mg L-1 18 mo after application. Their 18 Mg ha-1 application rate had a runoff P concentration of 1.17 mg L-1 that only decreased to 0.54 mg L-1 after 18 mo. Rostagno and Sosebee's (2001) data suggest that topical biannual or annual biosolids applications at 18 Mg ha-1 could maintain runoff P concentrations of >0.5 mg L-1.

The calcareous soils at the MERCO site might act to mitigate P levels in the runoff. Sharpley et al. (1981) observed that soluble P concentrations in runoff were inversely related to sediment concentrations and concluded that soil materials can act to reduce soluble P in runoff. Nevertheless, with such a small annual rainfall at the application site, soluble P would be expected to persist in the biosolids for a long time after application.

The presence of sparingly soluble phosphates in the biosolids explains why Rostagno and Sosebee (2001) showed little decrease in runoff water P concentrations 18 mo after application. Lime-stabilized biosolids rather than PSRP biosolids would be expected to yield lower runoff P concentrations. The high pH and high Ca concentrations produced by CaO treatment should precipitate Ca phosphates, such as octocal, from the more soluble phosphates in the biosolids. Likewise, Soon and Bates (1982) concluded that soluble P concentrations in soils amended with Ca(OH)2–treated biosolids were controlled by octocalcium phosphate. Fish et al. (2000), however, measured no differences between the P concentrations in the H2O leachates of PSRP and lime-stabilized biosolids collected after 5 d. Yet, Westermann (1992) reported that lime (CaCO3) additions to soils decreased available P contents and plant uptake of P.

A sequential extraction or selective dissolution was used to identify the forms of P in the biosolids (Table 3). The NaOH treatment removed almost all of the P from meta-strengite and variscite. In contrast, the HCl treatment dissolved most of the P in fluorapatite. Most of the P in the biosolids was extracted by NaOH, which suggests the presence of Fe- and/or Al-phosphates. Based on this extraction, 66 to 78% of the inorganic P in the biosolids was present as Fe or Al phosphates. Iron oxides and oxyhydroxides were probably contributed to the wastewater by local soils and rust from pipes and other sources. Biosolids often contain phosphates similar to those found in heavily fertilized soils. Lentner (1981) noted that most of the P in human feces is present as calcium phosphates. A great variety of orthophosphates, such as monocal and dical, have been added to human foods (Ellinger, 1983). Hence, the NaOH solubility of phosphates, such as dical, confounded the interpretation of the NaOH extract (Table 3). The citrate–bicarbonate–dithionite extracts of the biosolids contained 16 to 21% of total inorganic P, which suggests P occluded in Fe oxides. This interpretation, however, is also confounded by the likely presence of trical. The HCl-soluble fraction of the biosolids ranged from 5 to 12% of total inorganic P and was consistent with the presence of apatite and/or trical.


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  		Table 3. Sequential P extraction or selective dissolution to distinguish types of phosphates.{dagger}

 
Inorganic Analysis
The chemical compositions of the biosolids ash samples with greater exposure ages had greater contents of SiO2, Fe2O3, PbO, Cr2O3, and HgO (Table 4). In contrast, the ash from biosolids with greater exposure ages had lower concentrations of P2O5, MgO, CaO, and ZnO. These trends suggest that SiO2, Fe2O3, PbO, Cr2O3, and HgO contents have increased due to the loss of P2O5, MgO, CaO, and ZnO. The metals concentration in New York City biosolids, however, decreased from 1992 to 1998 (Department of Environmental Protection, 1998c) and the older biosolids might have contained somewhat higher concentrations of Pb, Cr, and Hg. The concentrations of As and Cd were below detection limits.


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  		Table 4. Elemental analysis of biosolids ash.{dagger}

 
Zartman (1993) reported Pb, Zn, and Cu concentrations of 13.1, 40.7, and 11.5 mg kg-1 for soils at the MERCO site before biosolids applications. These concentrations were at the low end of the ranges listed by Pierzynski et al. (2000) for normal trace element concentrations in soils. The Pb, Zn, and Cu contents of the biosolids ash samples were about 1000 times greater than the initial compositions of soils at the MERCO site. A hectare-furrow-slice (i.e., soil in one hectare to a depth of 17.8 cm) contains about 2.24 Gg of soil. If mixed with the topsoil, however, an 18 Mg biosolids ha-1 application would only contribute Pb, Zn, and Cu concentrations of 2.5, 9.3, and 6.3 mg kg-1. More than five biosolids applications would be needed to double the Pb content. Without surface mixing, 10 or more applications would be needed to attain the metal-rich concentrations of 10000 mg kg-1 Pb and Zn listed by Pierzynski et al. (2000). Yet, only one application would be needed to reach the metal-rich Cu concentration of 2000 mg kg-1. The major source of Cu and Zn in New York City wastewater was from multifamily and single-family housing that was probably derived from Cu pipes and Zn-galvanized wastewater plumbing (Department of Environmental Protection, 1998c).

The ratios of Zn, Cu, Pb, Cr, and Hg oxides relative to SiO2 + Al2O3 + Fe2O3 in the biosolids ash samples were plotted versus exposure age (Fig. 2) . The SiO2, Al2O3, and Fe2O3 were the major inorganic constituents in the biosolids samples. Based on X-ray diffraction, most SiO2 and Al2O3 in the biosolids occurred as quartz and feldspars. The Fe2O3 might occur as Fe oxides and phosphates. These constituents are insoluble and immobile in the slightly alkaline environment at the site; no losses of Si, Al, or Fe would be expected. The plots for PbO, Cr2O3, and HgO are nearly horizontal, suggesting that no losses of these metals occurred (Fig. 2). Phosphorus in the biosolids might have acted to reduce Pb mobility. Hettiarachchi et al. (2001) observed that additions of P to Pb-contaminated soils significantly reduced bioavailable Pb, which they attributed to the precipitation of lead phosphates. In contrast to PbO, the ZnO and CuO ratios decreased with greater exposure age, suggesting losses of Zn and Cu (Fig. 2). Similarly, Rostagno and Sosebee (2001) reported greater Cu concentrations in runoff from plots treated with biosolids and that the Cu concentrations in the runoff increased with application rate. They also reported that Cu concentrations exceeded the upper limit (0.5 mg L-1) for livestock drinking water 15 d after biosolids were applied using the highest (90 Mg ha-1) application rate. Wester et al. (2000) observed that Cu concentrations in soil samples from biosolids plots treated in 1993 and sampled in 1998 increased at the 0- to 5-cm sampling depth, but not at depths of 5 to 25 cm. Yet, they noted no differences in soil Zn, Ni, Pb, or Cd concentrations between control plots and plots treated with biosolids. Graphs of CaO, MgO, and P2O5 ratios (not shown) suggested even greater losses with exposure. Similarly, Rostagno and Sosebee (2001) reported that P concentrations in runoff increased with application rate and decreased with exposure age.



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  		Fig. 2. Contents of ZnO, CuO, PbO, Cr2O3, and HgO in biosolids ash samples relative to major components (SiO2 + Al2O3 + Fe2O3).

 

   CONCLUSIONS
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
REFERENCES
 
Exposure of biosolids on the soil surface resulted in decomposition of applied organic matter and loss of associated nitrogen and sulfur. Decomposition and loss of organic matter concentrated the inorganic materials in the residue. The available P and water-soluble P contents were progressively less in the biosolids with greater exposure age. Rain has either leached this P into the soil or moved it away in runoff. Yet after 82 mo exposure on the soil surface, available P concentrations in the biosolids were still 20 times greater than the sufficiency level for plant growth. However, if mixed with an hectare-furrow-slice of soil, a 90 Mg ha-1 biosolids application rate would be needed to yield available P concentrations at the sufficiency level. Sequential extraction indicated that the biosolids contained as much as 66 to 78% of P as iron phosphates, 16 to 21% as Fe oxides, and 5 to 12% as insoluble calcium phosphates (e.g., apatite). Iron phosphates probably formed during anaerobic digestion at the wastewater treatment plants. Successive water extractions also indicated that fresh biosolids contain phosphates as soluble or more soluble than dical. After exposure on the soil surface for 59 mo or more, dissolution and removal of phosphates resulted in soluble P concentrations consistent with the presence of trical. Yet, even trical can support P concentrations that are greater than that necessary for algal growth in surface waters and biannual or annual biosolids applications at the MERCO site would act to maintain even higher P concentrations in runoff. Elemental ratios suggest that the forms of Pb, Cr, and Hg in the biosolids were insoluble and immobile and that these metals have not migrated. Elemental ratios, however, suggested that Zn and Cu have either leached into the soil or been carried away in runoff.


   NOTES
TOP
 NOTES
ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Publication T-4-514 from the College of Agricultural Sciences and Natural Resources funded by MERCO.


   REFERENCES
TOP
 NOTES
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 


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