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Occurrence and Fate of Pharmaceuticals and Personal Care Products (PPCPs) in Biosolids

^a Department of Crop & Soil Sciences, 3111 Miller Plant Sciences Building, The University of Georgia, Athens, GA 30602
^b Department of Civil Engineering, Kansas State University, Manhattan, KS 66506-2905
^c Department of Biological and Agricultural Engineering, The University of Georgia, Athens, GA 30602-1339

^* Corresponding author (kxia{at}uga.edu)

Received for publication February 29, 2004.

	ABSTRACT

TOP ABSTRACT INTRODUCTION PHARMACEUTICALS AND PERSONAL... FATE AND BEHAVIOR OF... ANALYTICAL METHODS FOR... FATE OF PHARMACEUTICALS AND... CONCLUSIONS REFERENCES

 
Each year, large quantities of pharmaceuticals and personal care products (PPCPs) are used worldwide. Once conveyed to wastewater treatment plants, PPCPs can remain unchanged or undergo partial or complete transformation during wastewater treatment processes before discharge into the environment via effluent and biosolids for land application. Biosolids can be a major sink for some PPCPs. Previous investigations have indicated that land application of biosolids may be a potential important route through which PPCPs enter the environment. However, no information is available on exactly how closely the concentrations of PPCPs in the environmental media are related to the land application of PPCP-containing biosolids. This paper reviews currently available information on the occurrence of PPCPs in biosolids, methods of analysis, the potential fate of PPCPs in biosolids-applied soils, and composting as a potential means for removal of PPCPs from biosolids.

Abbreviations: K_ow, octanol–water partition coefficient • NP, nonylphenol • PPCPs, pharmaceuticals and personal care products • WWTP, wastewater treatment plant

	INTRODUCTION

TOP ABSTRACT INTRODUCTION PHARMACEUTICALS AND PERSONAL... FATE AND BEHAVIOR OF... ANALYTICAL METHODS FOR... FATE OF PHARMACEUTICALS AND... CONCLUSIONS REFERENCES

 
OVER THE LAST 40 years, efforts to understand the occurrence, fate, and environmental effects of anthropogenic chemicals have largely focused on industrial compounds and agricultural pesticides. The emphasis was appropriate due to the chemicals' large production, concentrated usage, widespread occurrence in the environment at high levels, persistence, and acutely toxic or carcinogenic effects. More recently, chemicals representing active ingredients in PPCPs have emerged as environmental contaminants with potentially widespread environmental effects.

A wide range of PPCPs has been detected in a variety of environmental samples at levels ranging from ng kg^–1 up to g kg^–1 (Halling-Sørensen et al., 1998; Daughton and Ternes, 1999; Kolpin et al., 2002). With the development of sophisticated and sensitive analytical instruments, more and more PPCPs can be detected at trace levels in the environment. Some of the detected PPCPs in the environment exhibit negative hormonal and toxic effects on various organisms at concentrations as low as µg kg^–1 (Daughton and Ternes, 1999). However, the environmental effects of many other PPCPs are not known. The concentrations, fate, and environmental effects of some PPCPs have been discussed in detail in a series of review articles (Halling-Sørensen et al., 1998; Daughton and Ternes, 1999; Jørgensen and Halling-Sørensen, 2000; Daughton and Jones-Lepp, 2001; Snyder et al., 2003).

Different from industrial chemicals and agricultural pesticides, many PPCPs and their biological metabolites are bioactive and are introduced into the environment on a continual basis due to the worldwide frequent usage of PPCPs by humans or for animal husbandry (Daughton and Ternes, 1999). A variety of human use PPCPs are discharged into wastewater treatment plants (WWTPs) via excretion with urine and feces as parent compounds, conjugated compounds, or metabolites, and through washing or direct disposal. A variety of PPCPs have been detected at various concentrations in influents of WWTPs from different regions (Table 1). At WWTPs, PPCPs can remain unchanged or undergo transformation during the treatment processes before being discharged into the environment via effluent and biosolids. The pharmaceuticals administrated for animal husbandry and the biological transformation products are excreted with urine and feces. The compounds can eventually enter into the environment due to runoff and leaching from feed lots and waste lagoons at confined animal feeding operations or from agricultural land irrigated or fertilized with animal waste (Halling-Sørensen et al., 1998).

This paper reviews current information about the occurrence of PPCPs in biosolids, methods of analysis, the potential fate of PPCPs in biosolids-applied soils, and treatments that may reduce the levels of PPCPs in biosolids before introduction to the environment through land application.

	PHARMACEUTICALS AND PERSONAL CARE PRODUCTS OF CONCERN

TOP ABSTRACT INTRODUCTION PHARMACEUTICALS AND PERSONAL... FATE AND BEHAVIOR OF... ANALYTICAL METHODS FOR... FATE OF PHARMACEUTICALS AND... CONCLUSIONS REFERENCES

 
Each year, large quantities of pharmaceuticals are sold and consumed in the United States and worldwide for the diagnosis, treatment, alteration, or prevention of human diseases. From 1999 to 2002, pharmaceutical sales increased worldwide by about 25% to 424 billion U.S. dollars (German Association of Research-Based Pharmaceutical Companies, 2004). In 2002, about 51% of the worldwide pharmaceutical sales were in the United States and Canada, 25% were in Europe, 12% were in Japan, 8% were in Africa, Asia, and Australia (not including Japan), and 4% were in Latin America (European Federation of Pharmaceutical Industries and Associations, 2003). Table 2 lists some of the most widely used pharmaceuticals in the United States. About 60% of the pharmaceuticals used in the United States are over-the-counter nonprescription drugs (Madhavan, 1994), mostly for pain relief, cold and flu, allergy, etc. (Harris Interactive, 2002).

In addition to pharmaceutical compounds, large quantities of personal care products, such as food supplements, fragrances, skin care and hair care products, insect repellents, cleaning products, and flame retardants are produced and sold in large quantities worldwide each year (Table 3). Synthetic musks, for example, are a group of chemicals used as fragrances in many personal care products, such as soap, perfumes, detergents, shampoos, and other personal care products. Nitro musks and polycyclic musks are the most commonly used synthetic musks, and worldwide production in 1996 was 7600 Mg (Rimkus et al., 1999; Rimkus, 1999). One of the polycyclic musks, galaxolide (1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta[g]-2-benzopyran, HHCB), is included in the USEPA's list of high production volume (HPV) chemicals (those that are produced in or imported into the United States at 450 Mg yr^–1). Similarly, nonionic surfactants such as alkylphenol polyethoxylates (APnEOs) are widely used in many personal care products (Thiele et al., 1997). More than 200000 Mg yr^–1 APnEOs, almost half of the world's annual production, were produced in the United States (United States International Trade Commission, 1995).

Effluents of WWTPs are the most direct sources of PPCP contaminants to waterways (Halling-Sørensen et al., 1998; Calamari et al., 2003; Kolpin et al., 2004). However, PPCPs associated with land-applied biosolids can mobilize in soil and leach to ground water or enter surface water through runoff (Jjemba, 2002; Golet et al., 2003; Yang and Carlson, 2003; Pedersen et al., 2003).

	FATE AND BEHAVIOR OF PHARMACEUTICALS AND PERSONAL CARE PRODUCTS DURING WASTEWATER TREATMENT PROCESSES

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The average daily quantity of wastewater generated per capita in the United States is about 450 L (120 gallons), and contains approximately 240 mg L^–1 suspended solids (Hammer and Hammer, 2001). More than 80% of the suspended solids is organic matter. Typically, the larger the organic input to the WWTP, the greater is the amount of sewage sludge produced at the facility. Sewage sludge is digested through biological, chemical, and physical processes before it is dewatered to produce biosolids. A typical WWTP produces about 240 kg of dry biosolids per million liters of wastewater treated (Metcalf and Eddy, 1991).

Transformation of Pharmaceuticals and Personal Care Products during Wastewater Treatment Processes
The transformation of PPCPs in WWTPs varies with compound physicochemical properties and wastewater treatment conditions. During the wastewater treatment processes, the parent PPCPs, conjugates, and metabolites may be (i) completely transformed to CO₂, (ii) partially transformed producing metabolites, or (iii) unchanged (Jørgensen and Halling-Sørensen, 2000). Few studies have traced the fate of PPCPs during wastewater treatment processes. Most published work consists of laboratory incubation studies using biofilm reactors and batch reactors with activated sludge at room temperature (20–25°C). Studies conducted in laboratories may over- or underestimate transformation rates of PPCPs in WWTPs because temperature conditions in many WWTPs may be higher or lower than that simulated in the laboratory (Metcalf and Eddy, 1991). In addition, many operating conditions in WWTPs are difficult to simulate in the laboratory.

Destruction of PPCPs in wastewater has been looked at only recently and the results have been variable. Discussion on four examples is presented below for ibuprofen, 17-ethinylestradiol, diatrizoate, and cyclophosphamide together with the breakdown products.

The pain reliever ibuprofen (Table 2) was detected in the influents of several Switzerland WWTPs at concentrations of 1 to 3.3 µg L^–1 (Buser et al., 1999). Its metabolites, hydroxyibuprofen (OH-Ibu) and carboxyibuprofen (CA-Ibu) were detected in the influents at concentrations about 1.5 times higher than that of the parent compounds. However, much smaller concentrations (6–105 ng L^–1) of carboxyhydratropic acid (CA-HA), another metabolite of ibuprofen, were detected in the same influent samples. Aerobic laboratory results of the WWTP influent mixed with activated sludge indicated a rapid disappearance of these compounds from the liquid phase, to <1 to 3% of the original levels within 8 h. The authors suggested that no sorption of the test compounds by the sludge occurred during the incubation, but no experimental evidence was provided for this assertion. Laboratory incubation studies using batch reactors with activated sludge and synthetic wastewater (Zwiener et al., 2002) indicated that <10% of ibuprofen was degraded to OH-Ibu, CA-HA, and CA-Ibu within 50 h. A greater ibuprofen concentration (177 µg L^–1) was used by Zwiener et al. (2002) than by Buser et al. (1999). The major metabolites of oxic and anoxic degradations of ibuprofen were OH-Ibu and CA-HA, respectively, but CA-Ibu occurred under both oxic and anoxic conditions. No further significant degradation of these three metabolites was observed (Zwiener et al., 2002). They suggested that remaining 90% of the ibuprofen that disappeared from the liquid phase of the reactors may have been sorbed by the sludge, partially degraded to other metabolites, or completely degraded. It was observed that the efficiency of elimination of ibuprofen from biofilm reactors decreased with increasing initial ibuprofen concentration.

The active ingredient of contraceptive pills, 17-ethinylestradiol, is eliminated from the human body as conjugates with glucuronic acid and sulfate (Ranny, 1977). Once discharged into WWTPs, these conjugates cleave fairly rapidly in contact with the microorganisms in activated sludge to release 17-ethinylestradiol, which is stable under aerobic conditions of an activated sludge process (Ternes et al., 1999a). This conceptualization may explain why the concentration of 17-ethinylestradiol in the effluent of a German WWTP was detected at a level almost twice as great as in the influent (Ternes et al., 1999b). Compared with other steroidal hormones, 17-ethinylestradiol degradation in an undiluted mixed sewage liquor was only 20% over 24 h, whereas 75% of 17ß-estradiol, a natural hormone, mineralized over 24 h (Layton et al., 2000).

Diatrizoate incubated with fresh activated sludge for 2 wk was not degraded, whereas 85% of iopromide was transformed into two metabolites within 54 h (Kalsch, 1999). Both diatrizoate and iopromide are widely used X-ray contrast media for detailed images of soft tissues in X-ray radiography. Cyclophosphamide, one of the most frequently used agents in cancer chemotherapy, exhibited poor degradability during 30-d batch incubation with activated sludge (Steger-Hartmann et al., 1997).

Large volumes of wastewater (up to hundreds million liters) in a typical WWTP must be treated every day and the average wastewater retention time (hydraulic retention time) in a WWTP varies from <1 h to a few days. Explaining the wastewater treatment process further is beyond the scope of the paper and standard texts in the field are available for review (e.g., Metcalf and Eddy, 1991; Hammer and Hammer, 2001). Hydraulic retention time frames are generally shorter than the degradation half-lives of many PPCPs that enter WWTPs (Halling-Sørensen et al., 1998), resulting in discharge of some relatively soluble PPCPs in effluent before degradation can occur.

Most studies on the fate of contaminants during WWTP processes have focused on the concentrations of target compounds in effluent discharged to the environment. A WWTP is considered to have efficient removal if there is a significant concentration reduction for the compound of interest in effluent compared with concentrations in influent (Simonich et al., 2002). This concentration reduction in effluent may be due to sorption of the target compounds to the solid phase, rather than degradation in the solution phase (Keller et al., 2003; Golet et al., 2003).

Occurrence of Pharmaceuticals and Personal Care Products in Biosolids
Compared with the hydraulic retention time of the liquid phase, the residence time of biological solids (mean cell residence time) in a WWTP system is much greater, ranging from a few days up to 30 d, due to repeated recycling of biological growths and extracted waste organic matter from one treatment reactor to another (Hammer and Hammer, 2001). However, even this residence time is shorter than the half-lives of some PPCPs (Halling-Sørensen et al., 1998). More importantly, during the wastewater treatment processes, PPCP molecules can move into microsites within the solid phase matrix (an "aging" process), and the biodegradability of organic compounds can be reduced significantly (Hatzinger and Alexander, 1995; Kelsey et al., 1997; Nam et al., 1998; Alexander, 2000). Sequestration into microsites of the solid phase can slow or stop microbial degradation if the molecules are inaccessible to microorganisms or extracellular enzymes. Diffusion of sequestered molecules out of the microsites in the solid phase can be extremely slow (Hatzinger and Alexander, 1997).

During wastewater treatment, PPCP and metabolite compounds partition into the solid phase depending on compound hydrophobicity, which is related to the compound octanol–water partition coefficient (K_ow). The greater the K_ow of a compound, the more hydrophobic it is. Dobbs et al. (1989) found that the sorption of certain chlorinated organic compounds on primary, mixed-liquor, and digested solids from municipal wastewater treatment plants correlated positively with their log K_ow, ranging from 1.26 to 5.48 (Fig. 1) .

View larger version (19K): [in this window] [in a new window]   Fig. 1. Experimental correlation between octanol–water partition coefficient (Kow) of some chlorinated organic compounds and their partition coefficients (Kp) for sewage sludge (based on weight of organic matter in the sludge solid phase). Vertical dotted lines indicate the boundary for sorption tendency of organic compounds on the sludge solid phase (data adopted from Dobbs et al., 1989; Rogers, 1996, with permission).

Although little research has been found on the sorption of PPCPs on the sludge solid phase during wastewater treatment processes, the relationship between PPCP sorption on the sludge solid phase and compound K_ow values can be expected to be similar as that indicated in Fig. 1. As shown in Tables 2 and 3, many PPCPs have K_ow values in the range of medium to high tendency for sorption on the sludge solid phase (Fig. 1). The sorption mechanisms proposed by Wang et al. (1993) for toxic organic compounds in the sludge solid phase should also apply to the PPCPs with similar K_ow values. Previous laboratory and full-scale wastewater treatment studies suggest that sorption to sewage sludge is the main removal process for some PPCPs from the wastewater stream (Kümmerer et al., 2000; Golet et al., 2003; Keller et al., 2003).

There is limited information on levels of PPCPs in biosolids. Fragrances have been detected at levels ranging from 1.5 to 147 µg kg^–1 (dry mass) in biosolids from the United States, Switzerland, and the Netherlands (Berset et al., 2000; Difrancesco et al., 2004). Nonylphenol polyexthoxylates (NPnEOs) and nonylphenol (NP), a metabolite of NPnEOs, have been detected at concentrations as high as 981 mg kg^–1 (dry mass) and 1380 mg kg^–1 (dry mass), respectively, in biosolids from many U.S. states (La Guardia et al., 2001; Keller et al., 2003; Xia and Pillar, 2003). Brominated diphenylethers, commonly used fire retardants, have been detected at 32 to 4890 µg kg^–1 (dry mass) in biosolids from several WWTPs in the United States and the Netherlands (Hale et al., 2001; De Boer et al., 2003). Golet et al. (2002) detected 1.4 to 2.4 mg kg^–1 (dry mass) of fluoroquinolone antibacterial agents in biosolids samples from Switzerland.

Approximately 6.9 million dry Mg of biosolids were generated in the United States in 1998, and production is estimated to reach 8.2 million dry Mg by 2010 (USEPA, 1999). Methods for disposal of biosolids include landfill, incineration, and land application. By 2010, 48% of biosolids are predicted to be land-applied (USEPA, 1999). Land application of biosolids, containing contaminants sorbed to biosolids, may concentrate in soil over time, providing a reservoir of pollutants that could eventually enter waterways through leaching and runoff. Hence, there is a need to evaluate the presence and environmental fate of PPCPs in biosolids before the occurrence of detrimental effects to the environment.

	ANALYTICAL METHODS FOR PHARMACEUTICALS AND PERSONAL CARE PRODUCTS IN ENVIRONMENTAL SAMPLES

TOP ABSTRACT INTRODUCTION PHARMACEUTICALS AND PERSONAL... FATE AND BEHAVIOR OF... ANALYTICAL METHODS FOR... FATE OF PHARMACEUTICALS AND... CONCLUSIONS REFERENCES

 
Numerous PPCPs have been detected at concentrations as low as in the ng L^–1 range in a variety of environmental samples including surface water, ground water, drinking water, and sediments samples. Several review papers on analytical methods for PPCPs in water and sediment samples have been published (Simonich et al., 2000; Golet et al., 2001; Öllers et al., 2001; Petrovi et al., 2001; Sacher et al., 2001; Ternes, 2001; Hyötyläinen and Hartonen, 2002; Patterson et al., 2002), but there is limited information on methods for PPCPs analysis in biosolids (La Guardia et al., 2001; Golet et al., 2002; Patterson et al., 2002). The published methods on water and sediment samples can be a good foundation for biosolids method development.

Figures 2 and 3 illustrate the general procedures for extraction, cleanup, concentration, and analysis for sediment and water samples. Among all the extraction methods for sediments, accelerated solvent extraction (ASE) is the most robust. It is operated under high pressure and at a temperature that can be programmed to the optimum temperature for extraction, resulting in efficient and rapid extractions using a small volume of solvent (Richter et al., 1996). In addition, a gradient of multiple solvents can be used for ASE to extract compounds with varying hydrophobicities. A combination of adsorbents is packed in an open column before an extractant is passed through for cleanup. The most commonly used packing materials are modified silica, Florisil, alumina, and different types of carbon. Once interferences from the matrix are removed through cleanup steps, the sample is normally dried down and redissolved in a smaller volume of solvent to increase the concentration of the target compounds. Compounds in the concentrated samples are then identified and quantified using a variety of instruments. Liquid chromatography–mass spectrometry (LC–MS, LC–MS–MS) and gas chromatography–mass spectrometry (GC–MS, GC–MS–MS) are the most powerful tools because of their identification capabilities and high sensitivities. More specific procedure requirements for selected PPCPs are listed in Table 4.

View larger version (37K): [in this window] [in a new window]   Fig. 2. General procedures for pharmaceuticals and personal care products (PPCPs) analysis in sediment samples. A summary of methods is cited in Table 4. GC, gas chromatography; HPLC, high performance liquid chromatography; IC, ion chromatography; LC, liquid chromatography; MS, mass spectrometry.

View larger version (31K): [in this window] [in a new window]   Fig. 3. General procedures for pharmaceuticals and personal care products (PPCPs) analysis in water samples. A summary of methods is cited in Table 4. GC, gas chromatography; HPLC, high performance liquid chromatography; IC, ion chromatography; LC, liquid chromatography; MS, mass spectrometry.

View this table: [in this window] [in a new window]   Table 4. Methods for selected pharmaceuticals and personal care products (PPCPs) analysis in environmental samples. The general procedures are outlined in Fig. 2 and 3.

	FATE OF PHARMACEUTICALS AND PERSONAL CARE PRODUCTS IN BIOSOLIDS AND BIOSOLIDS-AMENDED SOILS

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View larger version (18K): [in this window] [in a new window]   Fig. 4. Conceptual trend for the persistence of organic chemicals in biosolids-amended soils. Curve A is for compounds that are volatile, water soluble, or easily degraded; Curves B and C are for compounds with intermediate sorption potential on biosolids; Curve D is for compounds whose initial degradation is rate-limited because the microbial population has not acclimated to the compounds; and Curve E is for compounds that are nonvolatile, relatively water insoluble, and recalcitrant (Beck et al., 1996, with permission).

Difrancesco et al. (2004) observed rapid dissipation of 22 fragrance compounds 3 mo after being spiked in biosolids-amended soils. Complete dissipation for most of the tested compounds, except musk ketone and AHTN, was observed within the 1-yr die-away experiment, using soils with a wide range of textures and organic matter contents. A laboratory batch study conducted by Kreuzig et al. (2003) suggested that the abiotic transformation of diclofenac in a clayey silty soil and a silty sandy soil was dominated by the formation of non-extractable residues and that photo-induced degradation less relevant. Both soils were not amended with biosolids. Matscheko et al. (2002) investigated the levels of polybrominated diphenylethers (PBDEs) in lands that received biosolids application. Elevated levels of PBDEs were detected in all the biosolids-amended soils, compared with background levels. Twenty years after application of biosolids at 25 Mg ha^–1 yr^–1 for four consecutive years, the concentration of total PBDEs was 840 µg kg^–1, almost 8000 times the background level in the area. The concentrations of the tested compounds in the biosolids applied were not provided by the authors. Significant accumulation of PBDEs in earthworms of the sites was also observed. This study suggests that PBDEs are persistent in the soil environment, perhaps mainly due to the compound's high hydrophobicity (Table 3).

Insignificant mineralization (approximately 0.49–0.58%) of sarafloxacin, a fluoroquinolone antibiotic, was observed in a loam soil, a silt loam soil, and a sandy loam soil after 80 d incubation under aerobic conditions (Marengo et al., 1997). Biosolids were not applied to those soils. Marengo et al. (1997) concluded that the minimal mineralization may have been due to strong binding of sarafloxacin to the soil and the consequent nonavailability to soil microorganisms. Colucci and Topp (2001) observed rapid microbial-mediated dissipation of 17-ethinylestradiol under aerobic condition in soils varying widely in texture and properties. The 17-ethinylestradiol dissipated to below the detection limit (50 µg kg^–1) within 22 d of incubation. The 17-ethinylestradiol (log K_ow = 3.8) is much more hydrophobic than sarafloxacin (log K_ow = 0.49).

Tolls (2001) determined the sorption of a variety of veterinary pharmaceuticals with log K_ow ranging from –2.34 to 3.5 in soils without biosolids amendment. The sorption coefficients to soils varied from 0.2 to 6000 L kg^–1, demonstrating that these chemicals display a wide range of mobility. The author suggested that mechanisms in addition to hydrophobic portioning played a significant role in sorption of the tested compounds in soils. For the highly hydrophilic compounds, ion exchange, ion bridging at clay surfaces, surface complexation, and hydrogen bonding appeared to be involved in their sorption in soils. Inevitably, those reactions are largely affected by soil pH, soil mineral composition, and soil solution chemistry. Tolls (2001) proposed that instead of merely considering the contribution of hydrophobic partitioning to sorption in soils, a conceptually more complete representation of sorption of pharmaceutical compounds should be adopted.

There have been few investigations on the mobility and transport of PPCPs in soils, especially biosolids-amended soils. Rabølle and Spliid (2000) reported that the weakly adsorbing olaquindox (an antibiotic, log K_ow = 0.11) completely leached through soil columns, whereas the stronger-adsorbing antibiotic tylosin (log K_ow = 3.14) was retained in different depths depending on the soil properties. Boxall et al. (2002) observed that sulfachloropyridazine, an ionic antibiotic with low sorption coefficients (0.9–1.8 L kg^–1), was rapidly transported to surface waters after application to land. Thiele-Bruhn (2003) suggested that fast leaching through soils by macropore or preferential transport facilitated by dissolved soil colloids was the major transport process for strongly sorbed PPCPs. Surface runoff from lands may also contribute to widespread detection of PPCPs in the water environment. A variety of PPCPs were identified by Pedersen et al. (2003) in surface runoff from agricultural fields irrigated with disinfected tertiary recycled water or wastewater effluent-dominated stream water.

Removal of Pharmaceuticals and Personal Care Products in Biosolids through Composting
The safest approach to avoid potential detrimental effects of biosolids-associated PPCPs to the environment is to ensure that the compounds are adequately degraded before biosolids land application. Composting has been used as an effective means to degrade xenobiotic organic contaminants such as pesticides, polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), trinitrotoluene (TNT), and perchlorate (Williams and Keehan, 1993; Rao et al., 1995; Wallace et al., 1998; Büyüksönmez et al., 1999; Weed et al., 1999). Composting may accelerate the degradation of organic contaminants due to their exposure to high microbial diversity and activity (especially thermophilic organisms), abundant substrates, high temperature, changing pH, and successive shifts in aerobic and anaerobic conditions in microenvironments within a composting system (Büyüksönmez et al., 1999; Barker and Bryson, 2002).

Results from a recent investigation of 13 WWTPs in three U.S. states showed significantly lower concentrations of NP (65–100% less) in composted biosolids than in fresh biosolids (Fig. 5) . A pilot laboratory-controlled composting study (Fig. 6) provided further evidence that NP can be effectively degraded during composting. Biosolids initially containing approximately 450 mg kg^–1 NP was mixed with wood shavings at dry weight ratios of 43:57, 65:35, and 84:16. The mixtures were then incubated aerobically at 25, 45, and 65°C with the moisture content maintained at 65% for up to 70 d. A temperature between 45 and 65°C is ideal for thermophilic microorganisms, while 25°C is an ideal temperature for mesophilic microorganisms (Büyüksönmez et al., 1999). In all treatments, a rapid degradation of NP occurred within 15 d of incubation, with a maximum removal rate of 80%. Temperature significantly affected NP degradation at early stage of incubation. For compost with a biosolids to wood shavings ratio of 43:57, 6 d of incubation at 65°C resulted in a 76% NP reduction, whereas 41% of NP was degraded at 45°C during the same period. At 25°C, only 7 and 22% of NP was degraded after an 8-d incubation for treatments with biosolids to wood shavings ratios of 84:16 and 65:35, respectively. After incubation for 15 d, no significant difference was observed for NP degradation between composts treated at 45 and 65°C. For both treatments, approximately 80% of NP was degraded at Day 15 and approximately 92% of NP was degraded at Day 43. However, at 25°C, the NP degradation rate was much slower compared with the treatments at the two higher temperatures. By Day 15, between 50 and 60% of NP was degraded for composts treated at 25°C and an additional 55 d was required to degrade >90% NP. Lower biosolids to wood shaving ratios appear to favor higher NP degradation rates. High temperatures can significantly hasten NP degradation during composting, but composting for longer times at lower temperatures can yield similar results. The data suggest that composting can effectively remove certain PPCPs in biosolids. More studies are needed to monitor the effects of compositing on the degradation of other PPCPs in biosolids and to determine the most effective composting treatment parameters.

View larger version (51K): [in this window] [in a new window]   Fig. 5. Concentrations of nonylphenol (NP) in biosolids and composts from 13 wastewater treatment plants (WWTPs) in Georgia, South Carolina, and Kansas. The numbers indicate the WWTPs listed in the table below the graph. b, Biosolids; c, compost; c-3mo, biosolids composted for 3 months; c-12mo, biosolids composted for 12 months; nd, not detectable (Xia and Pillar, 2003).

View larger version (18K): [in this window] [in a new window]   Fig. 6. Removal of nonylphenol (NP) during aerobic composting of biosolids (BS) mixed with wood shavings (WS) at different ratios and different incubation temperatures.

	CONCLUSIONS

TOP ABSTRACT INTRODUCTION PHARMACEUTICALS AND PERSONAL... FATE AND BEHAVIOR OF... ANALYTICAL METHODS FOR... FATE OF PHARMACEUTICALS AND... CONCLUSIONS REFERENCES

Biosolids often serve as a sink for PPCPs and their partial metabolites that are not completely degraded during wastewater treatment processes. Research has shown a positive relationship between a compound's hydrophobicity and its sorption on biosolids. Although PPCPs, such as fragrances, flame retardants, surfactants, and their metabolites, have been detected in biosolids, there is limited information on the occurrence of many other PPCPs in biosolids. This lack of information is largely due to analytical limitations because of the complexity of the biosolids matrix. As sensitive analytical instruments such as liquid chromatography–mass spectrometry and gas chromatography–mass spectrometry become more available and widely used, more PPCPs are expected to be identified and quantified in biosolids.

The few published investigations on the mobility and transport of PPCPs in biosolids-amended soils suggest that land application of biosolids can be a major route through which some PPCPs enter the environment. However, no information is available on how closely the concentrations of PPCPs in the environmental media are related to the land application of PPCP-containing biosolids. Although the characteristics of environmental behavior of PPCPs may be unique because of their close association with biosolids, existing environmental behavior models for many well-studied industrial compounds and agricultural pesticides should be tested on PPCPs of interests. More importantly, the direct effect of land application of PPCP-containing biosolids on organisms in the environment has not been thoroughly investigated. Without detailed information on these issues, assessment of environmental consequence of land application of PPCP-containing biosolids is a difficult task.

To prevent PPCPs from entering the environment, there is an urgent need to document effective wastewater and biosolids treatment techniques. A thorough understanding of transformation mechanisms of PPCPs is the foundation for developing better treatments. Composting can effectively remediate many xenobiotic organic contaminants. Recent pilot-scale composting results suggest that nonylphenol, a metabolite of surfactants commonly used in detergents, can be effectively degraded during biosolids composting. Composting may be effective in removing many other PPCPs in biosolids as well.

	ACKNOWLEDGMENTS

 
The authors wish to thank Mr. Gene Weeks of the Department of Crop and Soil Sciences at the University of Georgia for sample collection. We appreciate the assistance from the staff members at the 13 WWTPs mentioned in this paper.

	REFERENCES

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