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Published in J. Environ. Qual. 33:1924-1929 (2004).
© ASA, CSSA, SSSA
677 S. Segoe Rd., Madison, WI 53711 USA

SHORT COMMUNICATIONS

Changes in the Nature of Sewage Sludge Organic Matter During a Twenty-One-Month Incubation

Ronald J. Smernika,*, Ian W. Olivera and Mike J. McLaughlinb

a Soil and Land Systems, School of Earth and Environmental Sciences, The University of Adelaide, Waite Campus, Urrbrae, South Australia 5064, Australia
b CSIRO Land and Water, PMB2, Glen Osmond South Australia 5064, Australia

* Corresponding author (ronald.smernik{at}adelaide.edu.au).

Received for publication December 10, 2003.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 Materials and Methods
 Results and Discussion
 Conclusions
 REFERENCES
 
Six sewage sludges from five sewage treatment plants in Australia were incubated for up to 21 months. Carbon losses at the end of the 21-mo incubation varied substantially. The remaining organic matter was isolated by treatment with hydrofluoric acid (HF) and characterized using a range of solid-state 13C nuclear magnetic resonance (NMR) spectroscopic techniques. By every measure (signal distribution in cross polarization [CP] and Bloch decay [BD] spectra, carbon NMR observability determined by spin counting, and the appearance of proton spin relaxation editing subspectra), the chemical composition of the residual organic matter appeared to be little different from that of the original sludges, even for those sludges that experienced the greatest carbon losses. Importantly, these NMR properties distinguish sewage sludge organic matter from soil organic matter. Thus, it should be possible to follow the decomposition of sewage sludge organic matter applied to soils in the field using solid-state 13C NMR spectroscopy.

Abbreviations: BD, Bloch decay • Cobs, proportion of potential carbon-13 NMR signal intensity detected in NMR spectrum • CP, cross polarization • NMR, nuclear magnetic resonance • PSRE, proton spin relaxation editing • T1H, proton spin–lattice relaxation rate in the static frame


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
Materials and Methods
 Results and Discussion
 Conclusions
 REFERENCES
 
THE AMOUNT OF SEWAGE SLUDGE produced worldwide is ever-increasing and a growing proportion is being applied to agricultural land (Düring and Gäth, 2002). Sewage sludge application provides potential benefits for agricultural soils including nitrogen (Binder et al., 2002) and phosphorus (Hogan et al., 2001) fertilization and increased organic carbon contents (Nyamangara and Mzezewa, 2001). However, sewage sludge also often contains high concentrations of potentially toxic organic chemicals (Wang and Jones, 1994; Jones and Sewart, 1997) and heavy metals (Hooda and Alloway, 1994; McBride, 1995; Antoniadis and Alloway, 2002; Oudeh et al., 2002; Walter et al., 2002).

The bioavailability of these toxic species in sewage sludge is often limited by the high sorption capacity of the sludges for these species (Brown et al., 1998; Li et al., 2001; Northcott and Jones, 2001). However, the longevity of this protective mechanism has been questioned, especially if organic binding sites are involved, since the organic matter eventually will be mineralized (Hooda and Alloway, 1994; McBride, 1995). Rowell et al. (2001) reported 20 to 40% losses of organic matter in just over a year in incubations of four different sewage sludges under greenhouse and field conditions.

Transformations of sewage sludge organic matter other than complete mineralization may also affect its sorption properties. In this study, we used a range of solid-state 13C NMR spectroscopic techniques to gauge changes in the chemical composition of sewage sludge organic matter during aerobic incubation. This should provide a better basis for understanding and predicting how sorption properties of soils change in the years following addition of sewage sludge.


   Materials and Methods
TOP
 ABSTRACT
 INTRODUCTION
 Materials and Methods
Results and Discussion
 Conclusions
 REFERENCES
 
Sewage Sludge Incubation and Demineralization
The six sewage sludges were sourced from five Australian sewage treatment plants. Details of sludge sources and composition can be found in our earlier paper (Smernik et al., 2003a). Briefly, Bolivar 95 and Bolivar 97 sludges were collected from Bolivar Sewage Treatment Works, which processes the majority of the domestic and industrial wastewater for Adelaide, South Australia. The plant processes approximately 130 ML of wastewater per day. Sludge from primary and secondary (trickle filtration) treatment processes are combined, anaerobically digested, dried in evaporation basins, and finally stockpiled. At the time of collection, Bolivar 95 sludge was 5 yr old and Bolivar 97 sludge was 3 yr old. Chelsea 96 was collected from the Eastern Treatment Plant, one of two plants servicing metropolitan Melbourne, Victoria. The plant processes 350 to 400 ML wastewater per day, mostly of domestic origin. Sludges from primary and secondary (activated sludge) treatment are combined, anaerobically digested, dried evaporatively, and stockpiled. When collected, Chelsea 96 sludge was 3 yr old. Werribee 97 was collected from the Western Treatment Plant, Victoria. The plant services the city of Melbourne and its surrounding region, and is the largest plant in Australia in terms of area. The plant processes domestic wastewater and 80% of Melbourne's industrial waste (500 ML/d, total). Sewage is processed by land filtration and lagoon stabilization. Drying is by evaporation. West Hornsby was collected from West Hornsby Treatment Plant, New South Wales. The plant processes 10 ML wastewater per day, mostly of domestic origin. Port Kembla was taken from Port Kembla Treatment Plant, New South Wales. The plant processes wastewater (13 ML/d) from heavy industry (including metal smelters) and domestic sources. Sludge undergoes primary sedimentation and anaerobic digestion before dewatering. Port Kembla sludge was less than 5 yr old when collected.

Solid-state 13C cross polarization (CP) and Bloch decay (BD) NMR spectra can be found in our earlier paper (Smernik et al., 2003a). The sludges were further characterized by the novel solid-state 13C NMR techniques PSRE (proton spin relaxation editing) and RESTORE [Restoration of Spectra via TCH and T1{rho}H (T One Rho H) Editing], which were used to probe the submicron heterogeneity of the sludges and their organic matter domain structure (Smernik et al., 2003b).

Incubations were performed by placing 1.6 kg of each air-dried sludge in separate plastic buckets. Deionized water was added to raise the moisture content of each sludge to 60% of that at field capacity (determined by two days of drainage at 100-cm suction). The lid of each bucket was pierced four times with a large nail, to allow air flow, before being secured. The mass of each bucket was recorded, and all were placed inside a dark incubation chamber where the temperature was maintained at 25 to 30°C. Once per week moisture lost was replaced by adding deionized water (mass basis). Subsamples of the sludges were taken after 6 mo of incubation by removing half of the volume. These 6-mo samples were dried at 35°C. Once dried, the samples were ground to pass a 2-mm sieve and stored in sealed containers for testing. The moisture content of the dried material was determined on subsamples (by heating at 105°C for 48 h) and used to calculate the oven-dry equivalent mass tested in all subsequent analyses. After the 6-mo sample was removed, the portion remaining in each treatment bucket was returned to the incubation chamber and maintained under the set conditions for a further 15 mo. At the end of the incubation period, a total of 21 mo, the remaining portion of each treatment was collected and processed as per the procedure used at the 6-mo collection stage.

Subsamples of the sewage sludges incubated for 21 mo were treated with hydrofluoric acid (HF) to isolate the organic matter for NMR analysis, according to the method of Skjemstad et al. (1994). The NMR analysis was not performed on samples incubated for 6 mo. Note that great care must be taken when working with HF to ensure against contact with the skin or inhalation of vapors. Even relatively small exposures can result in serious injury or death. Carbon recoveries on HF treatment were in the range of 70 to 94%. These recoveries are comparable with values of 79 to 87% determined for the sludges before incubation (Smernik et al., 2003a).

Nuclear Magnetic Resonance Spectroscopy
Solid-state 13C magic angle spinning (MAS) NMR spectra were obtained at a 13C frequency of 50.3 MHz on a Varian (Palo Alto, CA) Unity200 spectrometer. Samples were packed in a 7-mm-diameter cylindrical zirconia rotor with Kel-F end-caps and spun at 5000 ± 100 Hz in a Doty Scientific (Columbia, SC) MAS probe. Free induction decays were acquired with a sweep width of 40 kHz; 1216 data points were collected over an acquisition time of 15 ms. All spectra were zero-filled to 8192 data points and processed with a 50-Hz Lorentzian line broadening and a 0.005-s Gaussian broadening. Chemical shifts were externally referenced to the methyl resonance of hexamethylbenzene at 17.36 ppm.

The CP spectra were acquired using a 1-ms contact time and a 4-s recycle delay; 4000 transients were collected for each spectrum. The BD spectra were acquired using a 6.0 µs (90°) 13C pulse. A recycle delay of 90 s was used for all samples; 1000 transients were collected for each sample. The BD spectra were corrected for background signal (Smernik and Oades, 2001). Inversion–recovery experiments (Smernik et al., 2000) were performed to determine the rate and uniformity of T1H (proton spin–lattice relaxation rate in the static frame) relaxation. Thirteen recovery delays of between 0.1 ms and 2 s were used. A 1-ms contact time and a 2-s recycle delay were used; 2000 transients were collected for each spectrum.

Spin counting experiments were performed using the method of Smernik and Oades (2000a)(2000b). Glycine (AR grade; Ajax Chemicals, Seven Hills, NSW, Australia) was used as an external intensity standard (i.e., the glycine spectrum was acquired separately to those of the samples). For CP spin counting experiments, differences in spin dynamics between the sample and the glycine standard were accounted for using the method of Smernik and Oades (2000a), except that a variable spin lock (VSL) rather than a variable contact time (VCT) experiment was used to determine the rate of 1H spin–lattice relaxation in the rotating frame (Smernik et al., 2002). The results of spin counting are presented in Table 1. The carbon NMR observabilities for CP spectra, Cobs (CP), ranged from 70 to 82% and the carbon NMR observabilities for BD spectra, Cobs (BD), ranged from 97 to 111%. These values are again comparable with the values of 67 to 75 and 92 to 100% for Cobs (CP) and Cobs (BD), respectively, reported for the sludges before incubation (Smernik et al., 2003a). Errors in carbon NMR observabilities (Cobs values) are estimated to be ±10% in Cobs (CP) and ±15% in Cobs (BD) (Smernik and Oades, 2000a). Inversion–recovery experiments were analyzed by statistically comparing one- and two-T1H component fits to the data, using the method of Smernik et al. (2000).


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  		Table 1. Carbon-13 nuclear magnetic resonance (NMR) observabilities (Cobs) of hydrofluoric acid (HF)-treated sludges after 21 mo of incubation, determined from spin counting experiments.{dagger}

 
Elemental Analyses
Carbon contents of both whole and HF-treated sludges were measured using a LECO (St. Joseph, MI) CR12 carbon analyzer.


   Results and Discussion
TOP
 ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results and Discussion
Conclusions
 REFERENCES
 
Carbon contents of the sludges before incubation and after 6 and 21 mo of incubation are shown in Fig. 1. In each case carbon content decreased, although the magnitude and the timing of the decrease varied considerably. For the two Bolivar sludges, moderate initial decreases at 6 mo were followed by further decreases at 21 mo. Port Kembla and West Hornsby showed the largest decreases at 6 mo but only small further decreases at 21 mo. Carbon losses for Chelsea 96 and Werribee 97 were relatively small at both 6 and 21 mo. Clearly, the apparent increase in carbon content between 6 and 21 mo for the Chelsea 96 sludge is not possible, and is a reflection of the heterogeneity of carbon contents of the sludges in general, and of the Chelsea 96 sludge in particular. The extent of organic matter degradation was not related to the initial carbon content of the sludges. Indeed, the sludges with the highest (Werribee 97) and lowest (Chelsea 96) initial carbon contents (Fig. 1) experienced similar proportional carbon losses.



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  		Fig. 1. Organic carbon contents of the sewage sludge before incubation and after 6 and 21 mo of incubation.

 
Solid-state 13C NMR spectroscopy was used to determine the nature of the organic matter remaining after 21 mo of incubation. As will be seen below, only minor changes in the nature of the organic matter were detected after 21 mo of incubation, so NMR analysis was not performed on the sludges incubated for 6 mo.

The CP and BD 13C NMR spectra of the six sludges incubated for 21 mo are shown in Fig. 2. The sludges were treated with HF before NMR analysis (Smernik et al., 2003a). The distribution of signal intensity between four chemical shift regions is shown in Table 2. The four chemical shift regions are: 190 to 165 ppm assigned to carbonyl carbon in carboxylic acids, esters, and amides; 165 to 110 ppm assigned to aryl carbon, including O-aryl (165 to 145 ppm) as well as C- and H-substituted aryl carbon (145 to 110 ppm); 110 to 45 ppm assigned primarily to O-substituted alkyl carbon in carbohydrates, but also including methoxyl carbon and N-substituted alkyl carbon in protein; and 45 to 0 ppm assigned to alkyl carbon. The spectra in Fig. 2 are remarkably similar to the corresponding spectra of the sludges before incubation (Smernik et al., 2003a). This similarity is illustrated in Fig. 3, which shows changes in the distribution of CP spectral intensity due to incubation. Previous studies in our laboratory have indicated the level of precision obtainable in replicate measurements of chemical shift distribution for organic matter samples is ±2% (Baldock and Smernik, 2002). For the Bolivar 95, Bolivar 97, and Port Kembla sludges, differences in each spectral region signal intensity pre- and post-incubation are considerably less than 2%, and therefore the incubation has not had a significant effect on the distribution of 13C CP NMR signal intensity for these sludges. The biggest differences were found for Chelsea 96, for which there was an apparent 4% increase in O-alkyl C and a corresponding decrease in alkyl C of more than 3%. However, this is the sludge for which carbon loss was minimal over the 21-mo incubation. Differences in the nature of the organic matter for this sludge therefore cannot be attributed to different degradation rates for different components, but rather reflect the natural heterogeneity of this sludge. For the Werribee 97 and West Hornsby sludges there is a small (2–3%) increase in alkyl carbon on incubation, mainly at the expense of aryl and carbonyl carbon. These increases are barely above the level of precision in these measurements.



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  		Fig. 2. Solid-state 13C cross polarization (CP) and Bloch decay (BD) nuclear magnetic resonance (NMR) spectra of the sewage sludges after 21 mo of incubation.

 

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  		Table 2. Percent of total nuclear magnetic resonance (NMR) signal contained in four chemical shift regions in 13C cross polarization (CP) and Bloch decay (BD) NMR spectra of sewage sludges incubated for 21 mo.{dagger}

 


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  		Fig. 3. Changes in the distribution of solid-state 13C cross polarization (CP) nuclear magnetic resonance (NMR) spectral intensity brought about by incubation. Values larger than 2% are significant.

 
The absence of large changes in the nature of sludge organic matter on incubation apparently contradicts an earlier study in which we reported that organic matter in fresh sludge differed substantially from that in sludge stockpiled for one year in the field (Merrington et al., 2003). However, the comparison in this previous study relied on the composition of the sludge produced by the treatment works being constant over time. The methodology used in this current study allows a more reliable comparison, as it does not rely on this assumption.

As was found for the sludges before incubation (Smernik et al., 2003a), the BD spectra of the sludges after incubation contained considerably more alkyl carbon (Fig. 2, Table 2) than did the corresponding CP spectra, indicating the presence of alkyl C with a high degree of molecular mobility (Hu et al., 2000; Smernik and Oades, 2000a, 2000b; Preston, 2001). The results of spin counting experiments (Table 1) show that the BD spectra are essentially quantitative [Cobs (BD) = approximately 100%], while only 70 to 82% of potential NMR signal was detected using CP technique. Again, the CP and BD carbon observabilities are similar to those reported for the sludges before incubation (Smernik et al., 2003a). The under-representation of alkyl C with a high degree of molecular mobility accounts for much of this difference for the sludges both before and after incubation.

Analysis of NMR relaxation rates can provide additional and complementary information to that provided by the NMR spectra themselves. In particular, T1H relaxation rates can be used to probe the submicron heterogeneity of organic matter. Relaxation and spectral (chemical shift or frequency) information can be combined using the PSRE technique, which enables the generation of NMR subspectra of organic matter components characterized by different T1H relaxation rates (Newman and Hemmingson, 1990; Newman, 1992; Preston and Newman, 1992, 1995; Smernik et al., 2000; Petsch et al., 2001).

The T1H relaxation behavior of the sludges before and after incubation is compared in Table 3. Average T1H relaxation rates were determined from a one-component fit to inversion–recovery data. The range of average T1H relaxation rates was similar for the sludges before (74–160 ms) and after (90–223 ms) incubation. However, differences in average T1H relaxation rate pre- and post-incubation were quite large for some individual sludges, with decreases in average T1H recorded for Chelsea 96 (160 to 109 ms) and Port Kembla (153 to 90 ms), while increases in average T1H were recorded for Werribee 97 (148 to 223 ms) and West Hornsby (74 to 141 ms). Average T1H relaxation rates for the two Bolivar sludges were not greatly different pre- and post-incubation. These different behaviors do not correlate with the proportion of carbon mineralized during the incubation.


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  		Table 3. Results of one- and two-component fits to inversion–recovery data for sewage sludges after 21 mo of incubation.

 
The two main determinants of T1H relaxation rates are the concentration and nature of species with unpaired electrons (Smernik and Oades, 1999, 2000c) and molecular motion. Decreases in T1H relaxation rates on incubation may be caused by increasing concentrations of paramagnetic cations as organic matter is lost through mineralization. The decrease in T1H for Port Kembla is consistent with this explanation. However, the T1H relaxation rate for West Hornsby, which lost a similar proportion of organic C as did Port Kembla, almost doubled on incubation (Table 3). It is likely that molecular mobility is a more important determinant of T1H relaxation rates for these samples. In particular, T1H relaxation rates are very sensitive to water content (Newman, 1992; Hatcher and Wilson, 1991), as small increases in water content can greatly increase molecular motions that induce efficient T1H relaxation. Although all of the sludges were freeze-dried, they were then stored in sample vials that were not rigorously air-tight, and while being packed into the rotors for NMR analysis were subjected to different-length exposures to the laboratory atmosphere, the humidity of which also may have varied.

As was the case for the sludges before incubation (Smernik et al., 2003b), the inversion–recovery data for each incubated sludge were better described by a two-component fit than a one-component fit. The results of the two-component fits are presented in Table 3. Proton spin relaxation editing was used to generate subspectra of the sludge components characterized by rapid and slow rates of proton relaxation. The PSRE subspectra are shown in Fig. 4, and integral region data derived from these subspectra are presented in Table 4. The rapidly relaxing subspectra appear to be rich in protein and alkyl carbon, whereas the slowly relaxing subspectra are rich in O-alkyl carbon. Once again, the PSRE subspectra of the incubated sludges are very similar to those of the sludges before incubation (Smernik et al., 2003b), with the rapidly relaxing subspectra exhibiting similar features to bacterial biomass, and the slowly relaxing subspectra exhibiting similar features to partly degraded plant residues. It is interesting to note that although the T1H relaxation rates varied substantially between sludges pre- and post-incubation (Table 3), the differences in T1H relaxation rate between the two PSRE components were preserved. A similar situation was reported by Newman (1992) for wood samples, where moistening increased T1H relaxation rates for both cellulose and lignin, but T1H for the lignin component was always shorter than T1H for the cellulose component.



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  		Fig. 4. Solid-state 13C proton spin relaxation editing (PSRE) nuclear magnetic resonance (NMR) subspectra of the sewage sludges after 21 mo of incubation.

 

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  		Table 4. Percent of total nuclear magnetic resonance (NMR) signal contained in four chemical shift regions in 13C proton spin relaxation editing (PSRE) NMR spectra of sewage sludges after 21 mo of incubation.{dagger}

 

   Conclusions
TOP
 ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results and Discussion
 Conclusions
REFERENCES
 
The six sewage sludges investigated showed markedly different rates of carbon loss during a 21-mo incubation. Remarkably, the nature of the organic matter appeared to change very little during the incubation, even for those sludges for which the highest carbon losses were recorded. Previously we had shown that the organic matter in these sludges is heterogeneous, consisting of spatially and chemically distinct domains of bacterial and plant residues (Smernik et al., 2003b). One may have anticipated that degradation rates for these distinct domains would be different, and that incubation would result in a shift of average organic matter chemistry toward the more resistant component. That no such change was seen indicates that either degradation rates are very similar for the two components, or that it is not the nature of the organic matter that controls degradation rates in these sludges. That the overall rates of organic matter degradation are so dissimilar between the sludges, despite similar organic matter chemistries, suggests the latter is true (i.e., it is the nature of the sludge inorganic matter), or at least some component of it, that controls the rate of organic matter mineralization. This may occur through interactions between the organic matter and mineral surfaces that protect the organic matter from microbial attack (Golchin et al., 1994, 1996), or it may occur through the toxic effect of species such as heavy metal ions on the microbial community (Dahlin et al., 1997), either decreasing microbial degrader populations or selectively inhibiting microbial functions vital to the degradation process. Further work is required to distinguish between these possibilities and identify the inorganic species responsible.

Another important implication of this study is that the features of sludge organic matter that distinguish it from soil organic matter do not appear to change on incubation and partial degradation. The original sludge organic matter was distinctive in that it contained a large amount of alkyl carbon that was "invisible" to the 13C CP NMR technique, but visible to the 13C BD NMR technique. Furthermore, the rapidly relaxing PSRE subspectra were distinctively alkyl-rich, whereas for soil organic matter, it is generally the slowly relaxing PSRE subspectrum that is rich in alkyl C (Preston and Newman, 1992, 1995; Smernik and Oades, 1999; Smernik et al., 2000). Therefore, it should be possible to distinguish sludge organic matter from native soil organic matter in soils treated with sewage sludge. This will be tested in the future by applying these techniques to soils that have been amended with sewage sludge in the field.


   ACKNOWLEDGMENTS
 
This work was, in part, funded by an Australian Research Council (ARC) grant.


   REFERENCES
TOP
 ABSTRACT
 INTRODUCTION
 Materials and Methods
 Results and Discussion
 Conclusions
 REFERENCES
 


Related articles in JEQ:

This Issue in Journal of Environmental Quality

JEQ 2004 33: 1589-1599. [Full Text]  




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