Environ. Sci. Technol. 1998, 32, 821-827
Testing Anaerobic Biodegradability of Polymers in a Laboratory-Scale Simulated Landfill BRENDA B. RESS,† PHILIP P. CALVERT,† CHARLES A. PETTIGREW,‡ AND M O R T O N A . B A R L A Z * ,† Department of Civil Engineering, Box 7908, North Carolina State University, Raleigh, North Carolina 27695-7908, and The Procter & Gamble Company, Winton Hill Technical Center, 6100 Center Hill Road, P2N56, Cincinnati, Ohio 45224
The objective of this research was to develop an improved technique for measurement of anaerobic biodegradability that more closely simulates decomposition in landfills. Tests were conducted in 2-L reactors that contained a mixture of 14C-labeled test material, fresh refuse, and decomposed refuse as a seed. The four materials evaluated included a pure cellulose, a lignified cellulose (LC), a citric acid cross-linked cellulose (x-C), and a polyacrylate absorbent gel material (AGM). Material biodegradability, as measured by production of 14CH4 and 14CO2, was 55.5, 25.7, 52.0, and 2.5% for purified cellulose, LC, x-C, and AGM, respectively. Total recovery of radiolabel, after measurement of residual label in the leachate and decomposed refuse, was 77.4, 95.2, 74.1, and 66.7%, respectively. The reactor system provided repeatable results and simulated the refuse decomposition cycle in 6 months. Tests demonstrated that the biodegradability of x-C, a newly developed material, was comparable to that of pure cellulose and greater than that of LC. Cellulose in both forms is typically buried in landfills.
Introduction Recent estimates indicate that 209 million ton of municipal solid waste (MSW) is generated annually in the United States and that approximately 61% of this waste is disposed of by burial in a sanitary landfill (1). Development of integrated solid waste management programs, which include recycling and in some cases waste-to-energy, have led to a decrease in the use of landfills. However, there is a limit to the types of waste that can be recycled, and combustion has not been the solid waste management alternative of choice for many communities. Thus, landfills will be a significant part of MSW management for the foreseeable future. A complex series of chemical and biological reactions begins with the burial of MSW in a landfill (2). Initially, aerobic bacteria deplete the oxygen entrained in the refuse, and large amounts of carbon dioxide are produced. Thereafter, anaerobic conditions govern refuse decomposition because refuse buried in approximately 3-m-thick layers prevents oxygen replenishment. Since there is only a limited supply of alternate electron acceptors such as nitrate and sulfate, methanogenic conditions typically prevail. Cellulose * Corresponding author phone: 919-515-7676; fax: 919-515-7908; e-mail:
[email protected]. † North Carolina State University. ‡ The Procter & Gamble Company. S0013-936X(97)00296-4 CCC: $15.00 Published on Web 02/07/1998
1998 American Chemical Society
and hemicellulose are the principal biodegradable components of MSW, and the relationship between methane production and cellulose and hemicellulose biodegradation has been documented (3). To design products to be more compatible with biologically based waste management technologies, many companies are developing materials reported to be biodegradable (4-6). Since a large fraction of these materials will be disposed of in landfills, strategies to assess anaerobic biodegradability under conditions more representative of a landfill are required. Current protocols for anaerobic biodegradability testing utilize liquid-based culture systems more representative of anaerobic digesters than landfills and inocula that are not derived from refuse (5, 7). The objective of this research was to improve these existing protocols so that test conditions would more closely represent a landfill, while still allowing for a closely controlled experiment. It is recognized that the large scale and heterogeneity of landfills preclude a perfect laboratory simulation. A 2-L simulated landfill reactor was used to measure the anaerobic biodegradability of four radiolabeled polymeric materials. Anaerobic biodegradability was characterized by the production of 14CH4 and 14CO2, the accumulation of organic intermediates in leachate, and the presence of residual radiolabel in the solids at the completion of refuse decomposition.
Materials and Methods Experimental Design. Four radiolabeled materials were tested, including positive and negative controls, and two test polymers. The positive control was cellulose, chosen because its biodegradation is well-documented under simulated landfill conditions (3) and in landfills (8). The negative control was polyacrylate absorbent gel material (AGM), a lightly crosslinked polyacrylate that undergoes less than 2% mineralization under simulated landfill conditions (9, 10). One test polymer was a lignified cellulose (LC) that was expected to be degradable, but to a lesser extent than pure cellulose because lignin is known to interfere with anaerobic cellulose decomposition (11-14). LC represents a paper containing mechanical pulp such as newsprint (15). Together, pure cellulose and LC represent the range of celluloses and biodegradabilities typical of MSW. The second test material was a citric acid cross-linked cellulose (x-C) for which there were no data on anaerobic biodegradation in a high solids system. Reactors were filled with a mixture of shredded residential refuse, a seed of well-decomposed refuse that served to minimize the lag time prior to the onset of methane production, and a small amount of radiolabeled test material. Reactors were operated under conditions designed to enhance decomposition, so that decomposition could be completed in approximately 6 months (16). These conditions included the use of shredded refuse, leachate recycling and neutralization, and incubation at about 38 °C. While these operational conditions deviate from traditional landfill practice, their major effect is likely to be an increase in the decomposition rate instead of a change in a material’s biodegradability. In the absence of such enhancement, the decomposition cycle would require several years to complete, making it difficult for data on compound biodegradability to influence the introduction of new materials. Biodegradation tests were conducted in triplicate except for the LC that was tested in duplicate because of a limited amount of radiolabeled material. At the completion of the VOL. 32, NO. 6, 1998 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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decomposition cycle, reactors were destructively sampled for measurement of residual radiolabel. Materials. Refuse was collected in residential areas of Raleigh, NC, and shredded (2 cm × 5 cm) prior to use, as described previously (17). The seed was excavated from a landfill known to be in an active state of methane production, and it had been used to initiate methane production in previous experiments (15, 18). Radiolabeled cellulose, prepared by the incubation of tobacco leaves with 14CO2, was commercially available (American Radiolabeled Chemicals, St. Louis, MO). As purchased, the specific activity of the cellulose was high but unknown. Thus, it was necessary to both dilute and characterize the [14C]cellulose prior to use. This was done by ball milling the [14C]cellulose (∼10 mg) with 10 g of microcrystalline cellulose for 7 days. To verify that the labeled and unlabeled celluloses were well mixed, three subsamples of the mixture were hydrolyzed (72% H2SO4) followed by a secondary hydrolysis (3% H2SO4) (19). This resulted in complete cellulose dissolution. The hydrolysate was then analyzed by scintillation counting. Reproducibility among the three subsamples was good [coefficient of variation (CV) ) 0.23%], indicating that ball milling successfully mixed the two celluloses. The specific activity of the mixed cellulose was measured to be 227.3 µCi/g, and an average of 71.17 µCi (70.86-71.50) was added to each cellulose reactor. AGM had a specific activity of 2000 µCi/g, and an average of 72.68 µCi (71.59-74.26) was added to each reactor. Its synthesis was described previously (9). The LC was prepared by incubating [14C]glucose with pine twigs (Pinus strobus) followed by extraction of the nonlignocellulosic components (20). It had a specific activity of 1.39 µCi/g, and an average of 2.33 µCi (2.30-2.36) was added to each reactor. The preparation of radiolabeled x-C, in which individual cellulosic fibers have primarily intrafiber chemical cross-link bonds, has been described (21). The cross-linking agent was [14C]citric acid, the synthesis of which was also described previously (22). Prior to use, the x-C was washed 20 times in distilled water to remove the unreacted citric acid. After being washed, it had a specific activity of 21.79 µCi/g, and 71.34 µCi was added to each reactor. Equipment. Biodegradation experiments were conducted in 2-L reactors fabricated from two pieces (10.2-cmdiameter) of flanged, medium thickness glass tube (flat O-ring joint with 38.1 cm of glass below joint, Ace Glass, Vineland, NJ) (Figure 1). One end of each tube was sealed shut by a glass blower, and the opposite ends were connected with a Viton O-ring and a U-shaped flange clamp (Ace Glass) to form a gas-tight seal. A leachate collection vessel with a working volume of about 1 L was fabricated from a section of glass tubing not required for the reactor. Reactor inlets and outlets were constructed from glass stopcock fittings fused into the reactor and leachate collection vessel. A layer of fiber glass mesh was placed over the leachate drain to minimize clogging. The leachate collection and reactor vessels were connected with Kynar tubing (Cole Palmer, Niles, IL) and Swagelok (Crawford Fitting Co., Solon, OH) fittings. A small section of Pharmed tubing (Fisher Scientific, Pittsburgh, PA) was inserted between the Kynar tubing. This Pharmed tubing was inserted into a peristaltic pump to pump leachate to the top of the reactor. A magnetic stir bar was placed in the leachate vessel, and the vessel rested on a magnetic stirrer that was used to mix the leachate after buffer addition. All materials were selected to minimize both oxygen permeability and potential sorption of daughter products of the test materials to the reactor system. One connection from the top of the reactor led to a Tedlar gas bag (Pollution Measurement Corporation, Oak Park, IL). From the outlet glass stopcock, a Swagelok fitting and a Mininert valve were used to connect the reactor to the gas 822
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FIGURE 1. Reactor used for biodegradation test: A, reactor; B, leachate collection vessel with magnetic stirrer; C, gas collection bag; D, Kynar tubing (3/8 in. o.d., 1/16 in. wall thickness); E, peristaltic pump; F, Pharmed tubing (7/8 in. o.d.); G, hose pinch clamp; H, 6 mm Teflon stopcock with 3 in. glass extension; I, 3/8 in. tube to tube Swagelok union; J, Mininert valve in glass tube; K, 3/8 in. Swagelok union elbow; L, 3/8 in. Swagelok union tee; M, stainless steel flange clamp; N, hose clamp. bag. Gas bags were removed for gas volume measurement as described below. Reactor Loading. Reactors were filled with refuse that was mixed in a ratio of 30% seed to 70% fresh refuse by volume based on earlier work (15). An average of 492 dry g of seed and 499 dry g of fresh refuse was added to each reactor. Prior to loading, deionized water was added to the refuse to increase its moisture content to facilitate compaction. The test material was added to a reactor when it was half full and was then immediately covered by more refuse. Once full, layers of cheesecloth and 3-mm glass beads were placed over the refuse to facilitate distribution of recycled leachate. After sealing, sufficient water was added to produce about 500 mL of leachate. Reactors were tested for leaks by applying a vacuum and by monitoring the pressure for 5 min. The average density of refuse in the reactors was about 495 dry kg/m3 (835 wet kg/m3). Reactors were incubated at about 38 °C to enhance decomposition (16). Reactor Monitoring and Analytical Methods. Leachate was neutralized with 1 M KOH 5 or 6 days a week until the pH stabilized at about 7. Gas volume and concentration were measured semiweekly to biweekly depending on the gas production rate. CH4 and CO2 concentrations were measured by gas chromatography (17). To measure the 14CH4 and 14CO2 content of the gas, a 200-mL sample was injected in 50-mL increments into a system of two CO2 traps (2 M NaOH, 15 mL), followed by a tube furnace (Baxter Scientific, McGraw Park, IL) in which the CH4 was oxidized to CO2 at
700 °C, and then a second set of two CO2 traps. Samples were injected into the trapping system at 20 mL/min by using a syringe pump. Oxygen was used as the carrier gas at 30 mL/min. After trapping, 2 mL of each trap was transferred to Ultima Gold scintillation cocktail (Packard Instruments, Meriden, CT), refrigerated for 24 h to eliminate chemiluminescence, and then counted. Gas volume was measured in a specially fabricated stainless steel cylinder of known volume (3.461 L). The pressure in the cylinder was measured after evacuation and again after venting a gas bag into the cylinder. The amount of gas transferred from a gas bag to the cylinder was calculated by using the ideal gas law. It was typically necessary to repeat this procedure more than once to empty a gas bag. Methane production data are reported as dry gas at standard temperature and pressure. Dissolved 14CO2 and radiolabeled organics in the leachate were measured biweekly. After base addition, leachate samples were centrifuged and filtered (0.2 µm), and a 0.5mL subsample was added to 18 mL of Ultima Gold scintillation fluid for counting. Four milliliters of 0.2 M CdSO4 was added to 2 mL of centrifuged leachate to precipitate the 14CO2, after which the leachate sample was again filtered and counted. The counts remaining after precipitation were considered to be organic in origin, and 14CO2 was calculated as the difference between the filtered leachate and the precipitated leachate. In preliminary work, CdSO4 was shown to remove greater than 99.7% of the dissolved 14CO2. When gas production in the reactors approached zero, decomposition of the bulk refuse was nearly complete. However, it was possible that the test materials were still being converted to 14CH4 and 14CO2, but that this gas was not flushed into a gas bag by bulk gas production. To evaluate this, reactors were sparged with 2.5 L of N2 on two occasions, 35 days apart. Using this procedure, it was determined that radiolabeled gas production over the last 35 days of the experiment was less than 3% of total radiolabel production in all reactors. To terminate an experiment, reactors were destructively sampled, and solids were ground as described previously (18). Subsamples were analyzed for residual radiolabel in a Packard Sample Oxidizer and for cellulose, hemicellulose, and lignin (18). To evaluate whether selected test polymers were incorporated into cell mass, 15-g subsamples of ground refuse were sequentially extracted to produce fractions that represented protein and lipids (23). The presence of radiolabel in the protein and lipid fractions was compared with that in controls. Controls consisted of ground refuse containing freshly added test polymers in ratios designed to match the amount of radiolabel in the decomposed refuse.
Results and Discussion Refuse Decomposition. The average methane production rate for each reactor set is presented in Figure 2. This rate exhibited an exponential increase followed by an asymptotic decrease for all reactors as observed previously (2). Because a seed was used to initiate decomposition here, there was not a lag period prior to the onset of decomposition as had been reported in studies with no seed (2, 17). The methane yield for each reactor is presented in Table 1. All methane production was attributed to the fresh refuse because the methane potential of the seed was no higher than the value measured previously, 5.8 mL CH4/dry g (18), and its cellulose and hemicellulose concentrations were well below those in fresh refuse (Table 2). Both the methane yields and rates were similar across all reactors as would be expected for refuse from the same batch. Average cellulose, hemicellulose, and lignin concentrations for the refuse added to and removed from each reactor set are presented in Table 2. The cellulose and hemicellulose
FIGURE 2. Average methane production rate in each reactor set.
TABLE 1. Cumulative Methane Production in the Experimental Reactors reactor
cumulative methane (mL of CH4/g of fresh refuse)
cellulose-1 cellulose-2 cellulose-3 av (SD)
89.4 63.8 81.2 78.1 (13.1)
AGM-1 AGM-2 AGM-3 av (SD)
81.6 76.8 77.1 78.5 (2.7)
LC-1 LC-2 av (SD)
78.7 75.3 76.9 (2.4)
x-C-1 x-C-2 x-C-3 av (SD)
73.6 105.3 84.2 87.7 (16.1)
overall av (SD)
80.6 (10.5)
concentrations decreased over the decomposition period, but the lignin concentration increased since it is at best only slowly degradable under methanogenic conditions (11). An average of 80% of the cellulose and 67% of the hemicellulose present in the reactors initially degraded over the 182-day decomposition period (Table 2). This is consistent with previous reports in which 70-80% of the cellulose and 6077% of the hemicellulose decomposed (3, 17). Given the recalcitrance of lignin, a perfect mass balance would result in near 100% lignin recovery. The average recovery in Table 2, 92.9% (SD ) 13.6), is also consistent with previous reports. However, lignin recovery data must be reviewed with caution because the lignin concentration, as measured by the 72% sulfuric acid hydrolysis procedure used here, may change even if lignin is not degraded (24). The leachate pH was about 6 initially and increased rapidly to between 7 and 7.5 in all reactors (data not shown). The pH was externally neutralized for about the first 14 days of the experiment, after which neutralization was not necessary. The rapid pH increase can be attributed to the presence of seed that prevented an accumulation of carboxylic acids. In summary, the bulk refuse underwent methanogenic decomposition in a pattern consistent with previous studies and was consistent across all reactors. VOL. 32, NO. 6, 1998 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 2. Initial and Final Composition of Bulk Refuse in Experimental Reactors (% of Dry Weight) cellulose (%)
hemicellulose (%)
lignin (%)
fresh MSW seed weighted avc
48.2 6.0 27.3
10.6 2.4 6.5
Initial 14.5 19.4 16.9
cellulosed (SD) AGMd (SD) LCd (SD) x-Cd (SD) overall av (SD)
6.96 (3.5) 9.43 (1.4) 7.70 (0.3) 9.1 (1.3) 8.8 (1.1)
4.96 (2.6) 3.87 (0.45) 3.05 (0.1) 3.6 (0.67) 3.5 (0.5)
Final 22.4 (1.1) 29.8 (4.8) 22.9 (0.4) 26.1 (2.3) 25.5 (4.0)
cellulose loss (%)a
hemicellulose loss (%)a
lignin recovery (%)b
80.5 (1.9) 79.1 (2.6) 81.95 (1.2) 79.1 (2.9) 80.0 (2.3)
68.6 (3.5) 64.2 (3.8) 70.6 (1.6) 65.8 (6.1) 67.0 (4.4)
81.4 (4.4) 105.8 (17.8) 85.1 (0.2) 96.6 (8.0) 92.9 (13.6)
a Mass loss from weights of cellulose and hemicellulose added to and removed from each reactor. b Percentage of lignin added to each reactor that was recovered in the solids at the end of the experiment. c Calculated from the average mass of fresh MSW and seed added to each reactor. d Data are average of reactors in each set with the standard deviation (SD) given in parentheses.
FIGURE 3. Comparison of refuse and test material mineralization. Test Material Decomposition and Mass Balance. Decomposition of the radiolabeled test materials was monitored by the production of 14CH4(g), 14CO2(g), and 14CO2, and radiolabeled organics in the leachate. The average 14CH4 production rate for each reactor set is presented in Figure 3. As for bulk methane production, radiolabel methane production was repeatable (individual 14CH4 production rate plots not shown). The increase in 14CH4 production in the x-C reactors at the last time point indicates that some 14CH4 was sparged from the reactors with N2. Cumulative 14CH4 production as well as both gaseous and aqueous 14CO2 production for each reactor are summarized in Table 3. Test material mineralization was calculated from the total radiolabel recovered as gaseous 14CH4 and 14CO2 plus 824
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dissolved 14CO2. Average mineralization for the cellulose, AGM, LC, and x-C was 55.5%, 2.5%, 25.7%, and 52.0%, respectively (Table 3). These values do not include 14CO2 that may have precipitated as carbonates and been counted as residual solids in the combustion analysis. The accumulation of radiolabeled organic compounds in the leachate is also presented in Table 3. In the cellulose reactors, less than 0.3% of the initial label was present in the leachate at the end of the experiment, and soluble organics never exceeded 0.4% of the added [14C]cellulose (data not shown). The organic intermediates of cellulose production would include short-chain carboxylic acids that would not be expected to accumulate beyond steady-state pool levels in a stable methanogenic system. The 14C-labeled organic
TABLE 3. Mineralization and Mass Balance of Test Materials 14C-labeled
14C-labeled
(µCi)
mineralization (%)
organica (µCi)
solidsb (µCi) (SD)
radiolabel recovery (%)
20.08 15.42 19.50 18.33 2.54
0.46 0.74 0.73 0.64 0.16
60.29 47.75 58.39 55.47 6.76
0.20 0.21 0.23 0.21 0.02
14.31 (0.68) 17.57 (1.64) 14.39 (2.13) 15.42 1.86
80.77 72.62 78.95 77.44 4.28
0.99 0.91 0.84 0.91 0.08
0.65 0.63 0.51 0.60 0.08
0.25 0.31 0.31 0.29 0.03
2.64 2.49 2.30 2.48 0.17
0.07 0.09 0.08 0.08 0.01
32.89 (3.25) 46.20 (2.21) 60.74 (14.87) 46.61 13.93
48.68 64.83 86.57 66.69 19.01
LC-1 LC-2 av SD
0.33 0.28 0.31 0.04
0.31 0.25 0.28 0.04
0.02 0.01 0.02 0.01
27.82 23.52 25.67 3.04
ndc nd
1.30 (0.16) 1.93 (0.44) 1.62 0.45
83.02 107.44 95.23 17.27
x-C-1 x-C-2 x-C-3 av SD
14.50 14.79 14.29 14.53 0.25
17.80 19.67 18.12 18.53 1.0
4.10 5.18 4.53 4.60 0.54
51.03 53.32 51.78 52.04 1.17
2.09 1.99 1.91 2.00 0.09
13.21 (0.67) 14.56 (0.28) 14.08 (1.03) 13.95 0.69
72.46 75.58 74.2 74.08 1.56
reactor
cumulative 14CH 4 (µCi)
cumulative 14CO (g) 2 (µCi)
cellulose-1 cellulose-2 cellulose-3 av SD
22.18 17.98 21.31 20.49 2.22
AGM-1 AGM-2 AGM-3 av SD
14CO
2(aq)
a
a Dissolved 14CO or organics present in leachate at the end of the experiment. b Radiolabel in undegraded solids at the end of the experiment. 2 The standard deviation of triplicate analyses is given in parentheses. c nd, not detected.
fraction in the cellulose reactors may also have included soluble humic materials, although further characterization of the leachate was not performed. In the AGM reactors, about 0.1% of the added label was present in the leachate organic fraction at the end of the experiment, and this amount was nearly constant throughout (data not shown). As described below, AGM has a strong tendency to sorb, and this likely affected its minimal presence in leachate. In work with a different reactor system, 2.4% of the added AGM was present in the reactor leachate under simulated landfill conditions (9). It was hypothesized that the soluble material represented the low molecular weight components of the AGM. In the LC reactors, radiolabeled organics never exceeded 0.06% of the added radiolabel and were not detectable at the end of the experiment. Only the cellulose component of the LC was labeled. Thus, decomposition intermediates of the LC and cellulose would be similar. The lower amount of radiolabel in the LC reactors relative to the cellulose reactors is due to the lower amount of label added initially. In the x-C reactors, 2.8% of the added radiolabel was present in the leachate at the end of the experiment, and accumulation of soluble organics never exceeded 2.9% (data not shown). Mass balance recoveries were calculated by comparing the amount of label added to each reactor to the amount recovered as (a) gaseous end products plus (b) dissolved organic and inorganic label plus (c) residual radiolabel in the solids. The average recoveries were 77.4% (cellulose), 66.7% (AGM), 95.2% (LC), and 74.1% (x-C) (Table 3). Recoveries for the cellulose, LC, and x-C are within the range of typical experimental error and considered reasonable given the large number of measurements required to complete the mass balance. The radiolabel recovery in LC-2 above 100% suggests normal variation above the experimental, error-free value of 100%. Radiolabel recoveries in AGM-1 and AGM-2 were considerably lower than recoveries in other reactors. Less than 3% of the 14C-labeled AGM was accounted for as either gaseous end products or in the leachate. Thus, gas leakage is not a likely explanation for the low recovery, and 97% of the label should have remained in the residual refuse. Accurate analysis of the radiolabel present in the solids was dependent on good mixing. With the exception of AGM-3, the CV among the AGM reactors was comparable
to those in the other reactors (Table 3). Thus, the low mass balance recovery is not easily explained by poor mixing. However, Rittman et al. (25) reported that a low molecular weight fraction of AGM has a strong tendency to adsorb to glass and sand. The weighted average volatile solids content of the MSW and seed in the reactors was 51.7%; thus, it is likely that the AGM also sorbed to the refuse. It is possible that the AGM accumulated in a small fraction of the ground refuse and was not evenly distributed despite the low CVs in AGM-1 and -2. Trends in Test Material Decomposition. The average rates of refuse and test material mineralization are presented in Figure 3. In the cellulose reactors, there was excellent correspondence in the production rate trends for CH4, 14CH4, and 14CO2, indicating that the [14C]cellulose behaved in a manner similar to the refuse. Because the fresh refuse used in this research was 48.2% cellulose, similar behavior would be expected. In the LC reactors, there was close correspondence during the period of increasing CH4, 14CH4, and 14CO2 production. However, after the reactors reached their maximum rates, radiolabel production rates decreased more sharply, suggesting more rapid depletion of the bioavailable labeled cellulose. In the AGM reactors, the first and lower peak in radiolabel gas production corresponded with the maximum CH4 production rate, while the second and larger peak in radiolabel gas production was measured after the peak CH4 production rate. One explanation for this is that two fractions of AGM degraded. The degradability of the first fraction was comparable to that of refuse while that of the second fraction was slower. The potential for biodegradation of two distinct fractions of AGM is consistent with its large molecular weight distribution (25). An asymptotic decrease in mineralization rates was measured for both refuse and AGM. In the x-C reactors, the first peak in radiolabel gas production corresponded with the maximum CH4 production rate. The second peak in radiolabel gas production exceeded the first peak for 14CH4 but was lower for 14CO2. Essentially all of the unreacted citric acid was removed from the x-C prior to use. Thus, the first peak in radiolabel gas production rates cannot be explained by the availability of soluble citric VOL. 32, NO. 6, 1998 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 4. Presence of Radiolabel in Protein and Lipid Fractions of Decomposed Refusea
cellulose-2 control x-C-2 control
protein (SD) (dpm)
lipid (SD) (dpm)
159 853 (15 641) 66 993 (20 904) 33 609 (4 433) 11 300 (4 391)
194 928 (99 623) 142 209 (65 850) 27 612 (4 938) 3 272 (2 471)
a Data are the average of two and three 15-g subsamples for x-C-2 and cellulose-2, respectfully.
acid. However, the data suggest the presence of more than one fraction of x-C, with slightly different rates of biodegradation. As presented in Table 3, 55.5%, 2.5%, 25.7%, and 52.0% of the cellulose, AGM, LC, and x-C were mineralized, respectively. Relatively high mineralization of cellulose and low mineralization of AGM were expected based on previous reports (3, 9). AGM is a cross-linked polyacrylate, and it is estimated that 3-6% is not cross-linked and is water soluble (25). In previous research with AGM, approximately 0.5% of the AGM added to refuse reactors was mineralized, and it was hypothesized that the mineralized component was the low molecular weight fraction (9). Rittman et al. (26) measured approximately 2.2% mineralization of the soluble fraction of AGM in an anaerobic sand column. Mineralization was documented for the cellulosic component of LC. The LC decomposed to a lesser extent than cellulose as was expected given the presence of lignin (Table 3) (15). Mineralization of the x-C was also documented, and the extent of mineralization (52.0%) was not significantly different (p ) 0.05) from the pure cellulose (55.0%). Additional Fates of Test Materials in Biological Systems. The presence of radiolabel in the protein and lipid fractions from extracts of the decomposed solids in cellulose-2 and x-C-2 are presented in Table 4. The protein extracts for both reactors and the lipid extract from x-C-2 contained significantly more radiolabel (p ) 0.01) than the controls, suggesting that the cellulose and x-C had been incorporated into cell mass. To prove that the radiolabel in the protein extract was actually protein, the extract was acid hydrolyzed (27), after which the hydrolysate was passed through a cation exchange cartridge to immobilize amino acids. The amino acids were then eluted off the cartridge with 4 M NH4OH, the ammonia was evaporated, and the acids were counted by combustion. However, the results of this procedure were inconclusive. Other potential fates of materials that undergo anaerobic decomposition must also be recognized. Humification can be expected to occur during biodegradation, resulting in the incorporation of a parent compound into the humic fraction of a solid matrix. Finally, soluble degradation intermediates may form. Where appropriate, the toxicity of intermediates to methanogenic processes could be tested by using a modified anaerobic toxicity assay (18). Such intermediates may also become immobilized on the organic fraction of the refuse. Further work is required to develop techniques to distinguish cell mass, humic material, undegraded parent compound, and sorbed material in the residual solids. Summary. A system to measure the anaerobic biodegradability of a 14C-labeled test polymer under conditions representing the refuse matrix was developed in this research. The reproducibility across replicate reactors, the ability to obtain a mass balance on the added radiolabel, and the general trends in mineralization for the materials tested all suggest that the system is useful for measurement of biodegradation. In addition, the degradabilities of the positive and negative controls were consistent with previous research. 826
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Biodegradation in the laboratory-scale system described here will proceed more rapidly than what would be expected in a full-scale landfill because conditions here were optimized to accelerate decomposition. Refuse decomposition has been reported to occur in four phases (2), and these phases will occur in both a landfill and in a laboratory-scale simulation; only the time scale will vary. In this research, materials were tested in pure form as opposed to a form representative of their conditions upon post-consumer disposal. Thus, the biodegradabilities measured here were not reduced by inaccessibility of microbial enzymes to the substrate. Such inaccessibility might occur with a product made from several layers of different materials. Therefore, the biodegradability measured here should be taken as an upper limit of what could occur in a landfill. Nonetheless, our biodegradability test is more representative of a landfill than previously published test methods (5, 7) because of the use of a refuse matrix and refuse microorganisms. While we have demonstrated a system to measure material biodegradability, criteria have not been established to determine what extent of degradability should be required for a material to be termed “biodegradable under landfill conditions”. One possibility would be to use the results on cellulose and LC degradability measured here as benchmarks. Celluloses in a number of forms (office paper, boxes, newsprint, coated paper, packaging, etc.) are routinely buried in landfills, and cellulose degradation in landfills is welldocumented (8). In addition, Eleazer et al. (15) have documented the biodegradability of newsprint, office paper, coated paper, and corrugated containers under conditions similar to those used here. Thus, the data on radiolabeled cellulose and lignocellulose biodegradation measured here make it possible to compare the extent of degradability of a new material to that of materials routinely buried in landfills.
Acknowledgments Preliminary experiments conducted by Mr. John Eichenberger and Ms. Caroline Cline contributed to the techniques used for this study. The efforts of Dr. Brad Johnson to initiate this project, Ms. Barbara Nuck for the combustion analyses, and helpful comments on this manuscript by the anonymous reviewers and Dr. Tom Federle are gratefully acknowledged. We also thank Kathi McBlief for assistance with manuscript preparation. This research was supported by the Procter & Gamble Company, the Institute for Standards Research, and the National Science Foundation.
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Received for review March 31, 1997. Revised manuscript received November 20, 1997. Accepted December 17, 1997. ES970296H
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