Anaerobic Biodegradation of Aliphatic Polyesters - American

Biomacromolecules 2002, 3, 813-822

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Anaerobic Biodegradation of Aliphatic Polyesters: Poly(3-hydroxybutyrate-co-3-hydroxyoctanoate) and Poly(E-caprolactone) Thomas W. Federle,*,† Morton A. Barlaz,‡ Charles A. Pettigrew,† Kathy M. Kerr,† Joseph. J. Kemper,† Barbara A. Nuck,† and Lee A. Schechtman† The Procter & Gamble Company, P.O. Box 538707, Cincinnati, Ohio 45253-8707, and Department of Civil Engineering, North Carolina State University, Raleigh, North Carolina 27695-7908 Received February 7, 2002; Revised Manuscript Received May 5, 2002

Poly(3-hydroxybutyrate-co-3-hydroxyoctanoate), PHBO, represents a class of PHA copolymers that contain both short-chain-length and medium-chain-length repeat units. Radiolabeled and cold PHBO, containing 90 mol % 3-hydroxybutyrate and 10 mol % 3-hydroxyoctanoate were chemically synthesized using a new difunctional alkoxyzinc initiator. 14C-PHBO was incubated with samples of anaerobic digester sludge, septage, freshwater sediment, and marine sediment under conditions resembling those in situ. In addition, it was incubated in laboratory-scale landfill reactors. 14C-PCL (poly--caprolactone) was incubated with anaerobic digester sludge and in landfill reactors. Biodegradation was determined by measuring generation of 14CO2 and 14CH4 resulting from mineralization of the radiolabeled polymers. PHBO was extensively mineralized in digester sludge, septage sediments, and the landfill reactors, with half-lives less than 30 days. PCL was not significantly mineralized in digester sludge over 122 days. In the landfill reactors, PCL mineralization was slow and was preceded by a long lag period (>200 days), suggesting that PCL mineralization is limited by its rate of hydrolysis. The results indicate that PHBO is practically biodegradable in the major anaerobic habitats that it may enter. In contrast, anaerobic biodegradation of PCL is less ubiquitous and much slower. Introduction Recent estimates indicate that 217 million tons of municipal solid waste (MSW) is generated annually in the United States of which 9.9% is plastics.1 Approximately 55% of this waste is buried in sanitary landfills, and the remainder is managed through combustion (17%) and a combination of recycling and composting (28%). Integrated solid waste management programs that include recycling and occasionally combustion have resulted in a decrease in the use of landfills. However, there is a limit to the types of waste that can be recycled, and combustion is not the solid waste management alternative of choice for many communities. In the United States, only 5.2% of the 21.5 million tons of plastic waste generated annually is recycled. Thus, landfills will remain a significant part of MSW management for the foreseeable future. To create more environmentally compatible products, alternative plastics have been developed that are reported to be biodegradable.2 For many applications, biodegradation in anaerobic environments is of equal to or greater importance than biodegradation in aerobic habitats. Since a large fraction of plastic materials will be buried in a landfill, an assessment of anaerobic biodegradability under landfill conditions is required. In addition, biodegradable plastics could be incor* To whom correspondence may be addressed: phone, (513) 627-1166; fax, (513) 627-1208; e-mail, [email protected]. † The Procter & Gamble Company. ‡ North Carolina State University.

porated into flushable products, requiring an assessment of biodegradation in anaerobic digesters and septic tanks. Finally, plastics account for a significant fraction of the manmade debris found in oceans and waterways as a result of improper disposal, littering, or accidental release. Such debris can come to reside in anaerobic freshwater and marine sediments. Consequently, fate in a range of anaerobic environments is an important consideration for any biodegradable plastic. Poly(3-hydroxyalkanoates) (PHAs) are aliphatic polyesters produced in bacteria as carbon storage media. The most prevalent PHA found in nature is the homopolymer poly(3-hydroxyburyate), PHB. A variety of PHA copolymers are also known that fall in two main classes. The short-chainlength (scl) copolymer poly(3-hydroxybutyrate-co-3-hydroxyvalerate), PHBV, and the medium-chain-length (mcl) copolymers with repeat units made up of 3-hydroxyalkanoate units with six or more carbons, such as poly(3-hydroxyoctanoate).3 The natural polymers are perfectly isotactic with each stereocenter having the [R]-configuration, and they can have molecular weights over 1 000 000. PHBV has been commercially produced by fermentation under the name Biopol. Although currently expensive, predictions for the improvement of fermentation methodology with recombinant organisms or the production of PHA in crops suggest the cost could potentially decrease.4 Despite being considered a thermoplastic material, PHB undergoes significant thermal degradation in the melt and is a rather brittle material due to its high level of crystallinity.

10.1021/bm025520w CCC: $22.00 © 2002 American Chemical Society Published on Web 06/17/2002

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PHBV has somewhat better physical properties, a lower melt temperature (Tm) which varies somewhat based on valerate content and undergoes less thermal degradation in the melt.5 Recently, a third less common class of bacterial PHA copolymers has been found that consists primarily of scl units with some mcl comonomer units.6-9 Examples of this class of PHA copolymers are poly(3-hydroxybutyrate-co-3-hydroxyhexanoate) and PHBO. These copolymers have much better physical properties such as a lower Tm than PHBV at a given comonomer content and better extensibility.6,7 Thus, PHBO is representative of these less common PHA copolymers, which is currently being developed under the name Nodax. When this study commenced, natural PHBO was not readily available. Chemical synthesis facilitates control over copolymer composition and placement of the radiolabel. The chemical synthesis of PHA polymers and copolymers is typically accomplished by ring-opening polymerization of β-lactones as exemplified by the synthesis of PHBO (position of 14C radiolabel denoted by asterisk):

A new, soluble alkylzinc-diol initiator was used to prepare the PHBO. Although the literature contains numerous reports on initiators used to polymerize β-lactones,10-18 only the studies on aluminoxanes,10 distannoxanes,14 and alkylzinc alkoxide18 report the preparation of PHA with molecular weights >30 000. Since the desirable mechanical properties of PHA polymers are not realized until the weight-average molecular weights are >500 000,6 an initiator that produces high molecular weight, isotactic copolymer is needed. The new soluble alkylzinc alkoxide derived from 2,4-dimethyl2,4-pentanediol and diethylzinc meets these requirements.19 Multiple studies have examined the anaerobic biodegradability of PHB20-28 and various PHBV copolymers20,22,24-31 and have supported the anaerobic biodegradation potential of these polymers. Unlike PHB and PHBV, there are no reports in the literature demonstrating the anaerobic biodegradation of 3-hydroxyoctanoate containing polymers. Mergaert et al.25 investigated the degradation of mcl-PHA by incubation of a 1 mm thick film consisting of a copolymer containing 30% 3-hydroxyoctanoate and 70% 3-hydroxydecanoate in anaerobic sludge. After 123 days, this copolymer exhibited no significant weight loss, while injectionmolded 3 mm thick dogbone-shaped pieces of PHB, and PHBV copolymers containing 10% and 20% 3-hydroxyvalerate showed weight losses equal to 15, 7, and 11%, respectively, suggesting that an increase in monomer size may reduce the anaerobic biodegradation potential of PHA. Poly--caprolactone (PCL) is a synthetic aliphatic polyester with no stereocenters. PCL is generally regarded as a biodegradable material under aerobic conditions and has been documented to meet the American Society for Testing and Materials (ASTM) definition of biodegradability.32 For example, Benedict et al.33 tested PCL film biodegradation and reported bacterial and fungal cultures are capable of

Federle et al.

biodegrading and growing on PCL under aerobic conditions. Aerobic biodegradation tests conducted under realistic environmental conditions, using [U-14C]-PCL (U ) uniformly labeled), have confirmed the mineralization of PCL to CO2.34 However, studies examining the anaerobic biodegradability of PCL have yielded highly variable and equivocal results.24,28,29,31 An important consideration for a biodegradable polymer is that it biodegrades in all relevant environmental compartments, both aerobic and anaerobic. This paper examines the mineralization of radiolabeled PHBO and PCL to 14CO2 and 14CH under the anaerobic conditions found in sludge 4 digesters, septic tanks, sediments, and landfills by incubating these polymers with actual environmental samples from these habitats. In several of the studies 14C-(cellulose) lignocellulose was included as a reference. Experimental Section Materials for Preparation of PHBO. All reagents were dried or purchased in anhydrous form. Unlabeled [R]-βbutyrolactone (BL) was prepared by the chiral reduction of diketene 35 or by a multistep synthesis similar to a published procedure.36 14C-Labeled [R]-β-butyrolactone was prepared by this same multistep method starting with tert-butyl acetoacetate (American Radiolabeled Chemicals, Inc., St. Louis, MO) labeled at the keto-carbon which results in the methine of the lactone and subsequent polymer being the 14C-labeled position. Unlabeled monomer was distilled from CaH2, flash chromatographed down an alumina column with pentane as eluent, redistilled once or twice (fractional or spinning band column), and stored over activated alumina or activated molecular sieves for at least 24 h prior to use. Labeled monomer was diluted with unlabeled BL. It was then stirred over CaH2 and distilled (simple) under vacuum several times. The comonomer, 3-pentyl-β-propiolactone, was prepared and purified in a similar manner. Diethylzinc (1.1 M toluene solution from Aldrich, Milwaukee, WI) was used as received. Preparation of Initiator. A 0.1 M solution of diethylzinc was prepared by addition of a commercial 1.1 M solution of diethylzinc (9.9 mL) to dry toluene (100 mL) contained in a previously flame-dried, argon-filled flask. An equalmolar amount of argon-sparged, dry 2,4-dimethyl-2,4-pentanediol was slowly added to the diethylzinc solution over 5 min. An in-line bubbler was used to vent the ethane formed. After the addition was complete, the solution was allowed to stir for 20 min before use. The initiator solution was used the same day it was prepared. Preparation of PHBO. All manipulations of dry monomer and initiator were carried out in oven-dried, then flamedout, septa-capped glassware under dry argon with transfers made with a syringe or a cannula. Butyrolactone (labeled or unlabeled) was combined with 3-pentyl-β-propiolactone (9:1 molar ratio) and diluted with dry toluene to give a 15% monomer solution. The monomer solution was run through a short column of activated alumina directly into an argonfilled, flame-dried, reaction tube. The initiator (0.1 mol % based on monomer) was added via syringe, and the tube

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Anaerobic Biodegradation of PHBO and PCL Table 1. Characterization of PHBO Samples Prepared with Alkylzinc-diol Initiator

sample

Mn

Mw

Mw/Mn

comonomer content (mol %)

14C-PHBO

94 700 123 000

106 000 173 000

1.12 1.41

9.4 10.0

PHBO

Table 2. Characteristics of Sediments Used in Testing the Mineralization of PHBO under Anaerobic Conditions

specific activity (mCi/g) 0.169

heated by oil bath to 65 °C for 66 h to ensure complete reaction for the radiolabeled sample (typically 22 h is sufficient and was used for the cold sample). Afterward, a small aliquot of chloroform was added if needed to obtain a solution of reasonable viscosity, and the polymer was recovered by precipitation into an ether-hexane mixture (3:1 v/v). The solid was dried under vacuum (20-40 °C). Isolated yields of copolymers were 90-95%. Characterization of PHBO. Molecular weights and polydispersity were determined by size exclusion chromatography. For unlabeled copolymer, three Waters Ultrastyragel linear columns (one 50 × 7.8 mm and two 300 × 7.8 mm) in series were used with chloroform (1 mL/min) as the mobile phase. Labeled copolymer was analyzed with a bank of Shodex columns in series (one KF800 guard column, two KF804L, and one KF802) with THF as the mobile phase (1 mL/min). Equivalent Mn is obtained when a common unlabeled PHBO was run on either system. Calibration was performed with narrow molecular weight polystyrene standards, and the data were analyzed with Waters ExpertEase or Millennium software. The molecular weight data are given in Table 1. The comonomer content of the copolymers was determined by 1H NMR spectroscopy in CDCl3. The specific activity of the 14C-labeled PHBO was determined by combustion analysis with a Packard model 307 sample oxidizer to be 0.169 mCi/g. PHBO Test Samples. The dosing materials for the anaerobic mineralization tests in digester sludge, septage, and sediments were prepared by mixing 1 part radiolabeled PHBO with 9 parts nonradiolabled PHBO. Replicate subsamples were combusted to verify the specific activity. Preparation and Characterization of Lignocellulose. Radiolabeled natural lignocellulose was biosynthesized in Pinus strobus using the cut stem procedure described by Crawford and Crawford.37 This process involved feeding 14Cglucose to pine twigs, resulting in the preferential incorporation of 14C into 14C-(cellulose)-lignocellulose. Low molecular weight 14C-labeled contaminants were removed using the extraction procedures described by Crawford.38 The extracted lignocellulose was characterized by the Klason analysis for acid-soluble (cellulose) and acid-insoluble (lignin) 14C-labeled components.39 The final material had a specific activity of 1.39 µCi/g, and contained 29% acid insoluble and 70% acid soluble components. Preparation and Characterization of Poly--Caprolactone (PCL). Radiolabeled-PCL was synthesized using [U-14C]--caprolactone prepared from 14C-cyclohexanone (New England Nuclear) by a Baeyer-Villiger oxidation. The polymer was then prepared by bulk polymerization of a mixture of the labeled lactone and unlabeled caprolactone with a dialkylaluminum isopropoxide initiator in a manner

characteristic

marine sediment

freshwater sediment

pH organic matter (% dry wt) sand content (% dry wt) silt content (% dry wt) clay content (% dry wt)

7.8 4.7 75 17 8

6.7 3.6 79 17 4

similar to those described by Teyssie´ and co-workers.40 The [U-14C]-PCL was precipitated from methylene chloride into methanol, vacuum-dried, stored, and used as a fluffy powder. Combustion analysis of the powder indicated the [U-14C]PCL had a specific activity 0.975 mCi/g. Chromatographic analysis of the [U-14C]-PCL indicated that the Mn and Mw were 79 100 and 98 100, respectively, relative to polystyrene standards. The distribution of radioactivity in the various molecular weight fractions of the [U-14C]-PCL was associated with a single peak, indicating radiochemical purity. In addition, chromatographic analysis of the [U-14C]-PCL indicated that only background levels of radioactivity were detected before and after the main peak. Environmental Samples. Anaerobic digester sludge was collected from the Federalsburg Wastewater Treatment Plant in Federalsburg, MD, on the day of the test. This plant receives predominantly domestic wastewater. Sludge was collected in a polypropylene container, which was filled completely and tightly capped. This sample was immediately transported back to the laboratory and transferred into an anaerobic chamber. The total solids concentration was 29 917 mg/L at collection and was concentrated to 50 000 mg/L by settling. Equal volumes of concentrated sludge and mineral salts medium were combined as described previously.41 Septage was collected 1 day prior to testing from a functioning residential septic tank receiving toilet, kitchen, and laundry wastewater in Oxford, MD. Septage was collected through a riser using a tube immersed into the layer of solids located 2-3 m below the surface. Septage was placed into a polypropylene container, which was filled completely, tightly capped, and immediately transported back to the laboratory and placed in an anaerobic chamber. The septage was allowed to settle for 30 min, and the settled fraction was screened through a #8 mesh sieve. The pH of the septage was 6.2, and the redox potential measured with a platinum electrode was -301 mV. The levels of total suspended solids and volatile suspended solids were 29 300 and 16 200 mg/L, respectively. The septage was adjusted to approximately 8000 mg/L total suspended solids with anoxic water for use in the mineralization test. Marine and freshwater sediment with the overlying water were collected using a core tube. Marine sediment was collected from the Indian River Inlet in Delaware; freshwater sediment was collected from the Schuykill River in Valley Forge National Park, PA. Upon collection, the redox potentials of the marine and freshwater sediments were -345 and -244 mV, respectively. Additional characteristics of the sediments are shown in Table 2. The overlying water was decanted, and the sediments were stored in an anaerobic chamber for 3 or 5 days prior to use.

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Fresh refuse was collected from residential areas of Wake Co., NC, and shredded (2 cm × 5 cm) prior to use. Welldecomposed refuse was excavated from a landfill known to be in an active state of methane production, shredded, and then incubated at 37 °C prior to use. Mineralization in Digester Sludge, Septage, and Sediments. Mineralization of the test materials was determined using the batch test system described previously.41 Radiolabeled test chemicals were gravimetrically added to 500 mL test vessels containing 300 mL of digester sludge, septage, or sediment. Sediment was placed in the vessels in a manner that maintained the vertical integrity of the core. The test materials were mixed uniformly into the digester sludge and septage. In the case for sediments, the test materials were mixed into the top 1 cm layer. Final added test material concentrations in digester sludge were 10 mg/L for PCL and 100 mg/L for lignocellulose and PHBO. Final added concentrations of PHBO and lignocellulose in sediments were 50 mg/kg dry weight. Final added concentration of PHBO in septage was 100 mg/L. PHBO was tested in triplicate in digester sludge and in duplicate in septage. PCL was tested in triplicate in digester sludge. A single vessel containing the reference material, lignocellulose, was prepared for digester sludge and the sediments. Vessels containing digester sludge were incubated at 35 °C in a water bath, while those with septage and sediment were incubated at 22 °C. The headspace of each vessel was continuously purged with nitrogen gas. The effluent gas was passed through a series of potassium hydroxide (KOH) scrubbers to trap 14CO2. The effluent gas from the traps was mixed with oxygen and passed through a quartz combustion tube containing cupric oxide heated to 800 °C to convert 14CH4 to 14CO2, which was subsequently trapped in a second series of KOH traps. Production of 14CH4 and 14CO2 was monitored by measuring the level of radioactivity in the various traps by liquid scintillation counting (LSC). At the termination of each study, the vessels were acidified with 10 mL of 5 N sulfuric acid and incubated overnight to recover any dissolved 14CO2. With the septage and sediments, the contents of the vessels were dried at 105 °C and homogenized using a mortar and pestle. Five replicate subsamples were combusted using a Packard model 307 sample oxidizer, and radioactivity was determined by LSC. Kinetic Analysis of Mineralization Data. The average cumulative gas production data were fitted to various nonlinear first-order production equations as described by Federle and Nuck.41 using Jandel TableCurve 2D software. The most appropriate model was based upon r2, F-value, and visual inspection of the fit. Half-lives were estimated from the first-order rates Landfill Simulation. The methods utilized to measure polymer decomposition in a simulated landfill have been described previously42 and are summarized here. The 14Clabeled test polymers were added to laboratory-scale reactors filled with municipal solid waste (MSW). As the MSW decomposed, the production of 14C-endproducts and the presence of 14C in the reactor leachate were monitored. The polymers were tested in triplicate 2-L reactors that were filled with a mixture of shredded fresh MSW (523 g) and well-

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decomposed MSW (374 g) that served as an inoculum. The reactors were operated under conditions to accelerate decomposition, so that refuse decomposition could be completed in approximately 6 months. These conditions included the addition of well-decomposed refuse as an inoculum for rapid initiation of methanogenesis, the use of shredded MSW, daily recycling, and neutralization of leachate and an incubation temperature of 38 °C. An average of 10.27 mCi of 14C-labeled PHBO or 99 mCi of PCL was added to each reactor. The total volume and composition (CH4,14CH4, CO2, 14 CO2) of gas produced in each reactor were monitored. Gas was collected in Call-5-Bond gas sampling bags (Calibrated Instruments, Hawthorne, NY). Total volume was measured by transferring gas from the gas bag into an evacuated cylinder of known volume. Production of 14CH4 and 14CO2 was determined by flushing 400 mL of gas, in 50 mL increments, through a series of base traps and a tube furnace followed by a second series of traps. The amounts of 14CH4 and 14CO2 were determined by measuring the level of radioactivity in the various traps by LSC. The levels of total radioactivity and dissolved 14CO2 in the leachate were also measured. Subsamples of leachate were treated with base, filtered, and analyzed by LSC. Additional samples were treated with 2 M CdSO4 to precipitate dissolved 14CO2 and analyzed by LSC. The level of dissolved 14CO2 was determined by difference. At the completion of the decomposition cycle, the undecomposed refuse from the PHBO experiment was analyzed for residual radiolabel. The refuse was dried and ground to pass a 0.5 mm screen. Subsamples were then combusted in a series of two-tube furnaces designed to convert 14C-organics to 14CO2. The effluent from the furnaces was passed through base traps, which were analyzed by LSC. System operation was verified by combusting cellulose samples of known specific activity. The residual solids from the PCL tests were not analyzed. Abiotic Leaching Test. Leaching tests were conducted to evaluate the abiotic formation of soluble products from PCL and PHBO under landfill conditions. Tests were conducted using synthetic leachates at pH 5 and 7 to mimic the acid and methanogenic phases of decomposition, respectively. The former consisted of 8000 mg of acetate/L, 900 mg of propionate/L, and 6000 mg of butyrate/L in water, while the latter contained 100 mg/L acetate, 10 mg/L propionate, and 10 mg/L butyrate in water. Samples of radiolabeled PHBO and PCL were incubated with the synthetic leachates in triplicate 60-mL serum bottles at 38 °C under an anaerobic atmosphere. Biological activity was prevented by the addition of 250 mg of HgCl2/L. Radioactivity solubilized into the leachate was determined by LSC. Results Mineralization in Treatment Systems. In many sewage treatment plants, anaerobic digestion is used to reduce the volume of sludge. As an engineered system, sludge from anaerobic digesters serves as a reliable and robust source of anaerobic microorganisms capable of degrading an array of

Anaerobic Biodegradation of PHBO and PCL

Biomacromolecules, Vol. 3, No. 4, 2002 817 Table 3. Final Distribution (%) of Radioactivity following Incubation of Radiolabeled PHBO, PCL, and Lignocellulose in Digester Sludge, Septage, Marine Sediment, and Freshwater Sediment habitat test material digester sludge PHBO PCL lignocellulose septage PHBO marine sediment PHBO lignocellulose freshwater sediment PHBO lignocellulose

% 14CO2

% 14CH4

73.6 ( 13.1a 0.1 ( 0a 13.8b

14.2 ( 3.2a 0.1 ( 0a 8.6b

d d d

81.5 ( 0.1c

13.8 ( 1.0c

7.4 ( 1.3c

91.3 ( 0.7c 25.7b

1.3 ( 0.5c 9.6b

4.7 ( 0.9c 64.8b

88.4 ( 0.7c 23.4b

4.6 ( 0.8c 13.6b

7.0 ( 1.5c 63.1b

solids (%)

a Mean ( standard deviation (n ) 3). b (n ) 1). c Mean ( standard deviation (n ) 2). d Not determined.

Table 4. Rates and Extents of PCL, PHBO, and Lignocellulose Mineralization in Digester Sludge, Septage, Marine Sediment, and Freshwater Sediment habitat test material

Figure 1. Mineralization of polyhydroxybutyrate-octanoate (PHBO), lignocellulose, and polycaprolactone (PCL) to 14CO2 and 14CH4 in anaerobic digester sludge incubated at 35 °C (mean ( standard deviation). Test concentrations were 100 mg/L for PHBO and lignocellulose and 10 mg/L for PCL.

natural and synthetic materials. In actual practice, any plastics (e.g., flushable products) disposed down the drain would likely become a component of sludge and could be subjected to anaerobic digestion. The mineralization of PHBO and PCL to carbon dioxide and methane was examined in anaerobic sludge from a wastewater treatment plant receiving primarily household wastewater. Lignocellulose, which is natural and widespread, was included as a reference. Figure 1 shows the mineralization of these compounds in anaerobic digester sludge as a function of time. No significant mineralization of radiolabeled PCL was evident. Analysis of the aqueous phase at the end of the experiment indicated that only 0.2% of the dosed radioactivity was soluble, indicating that little PCL had hydrolyzed. In contrast, PHBO was rapidly and extensively mineralized to CO2 and CH4. Approximately 88% of the dosed radioactivity was recovered as 14CO2 and 14CH4 at the end of the experiment (Table 3). The estimated halflife of PHBO was 8 days, which is short compared to the residence time of most digesters, which is typically 30-60 days (Table 4). Lignocellulose also was mineralized but less extensively. Only 22.4% of radiolabeled lignocellulose was recovered as 14CO2 and 14CH4 after 61 days.

digester sludge PHBO PCL lignocellulose septage PHBO marine sediment PHBO lignocellulose freshwater sediment PHBO lignocellulose

duration total % (days) mineralized

first-order rate (days-1)

av half-life (days)

61 122 61

88 ( 16a 0.08 ( 0.01b 0.2 ( 0. 1a 0 22c 0.04 ( 0.01b

141

95 ( 1d

0.03 ( 0.00b

25.7

141 141

93 ( 1d 35c

0.12 + 0.01b 0.07 ( 0.01b

5.7 e

141 141

93 ( 1d 37c

0.13 ( 0.02b 0.09 ( 0.02b

5.4 e

8.2 e e

a Mean ( standard deviation (n ) 3). b Mean ( standard error. c (n ) 1). d Mean ( standard deviation (n ) 2). e Not determined due to incomplete mineralization.

In many regions, households are not connected to sewage treatment plants but depend on on-site treatment. Septic tanks are the most common form of such treatment. In such systems, any flushed objects would likely be retained in the tank, which is anaerobic. The mineralization of PHBO, which has a density >1, was examined in septage from a residential septic tank (Figure 2). PHBO was extensively mineralized (>95%) after 141 days (Table 3). The estimated half-life of PHBO in the septic tank was approximately 26 days (Table 4), which is short compared to typical solids residence times for most tanks, which is determined by the frequency of pumping, which rarely exceeds once every 2 years. Mineralization in Sediments. In aquatic environments, plastics may sink and become buried in the sediments. In organic-rich waters, sediments are usually anaerobic all year or during the summer, when metabolism is at its peak and oxygen saturation is at its lowest. Marine and freshwater sediments differ in metabolism. Marine waters are characterized by high levels of sulfate, while sulfate levels are typically low in freshwaters. In marine sediments, sulfate

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Federle et al.

Figure 2. Mineralization of 100 mg/L polyhydroxybutyrate-octanoate (PHBO) to 14CO2 and 14CH4 in septage from a residential septic tank incubated at 22 °C (mean ( standard deviation).

usually serves as the major terminal electron acceptor and sulfate reducing bacteria proliferate. In freshwater sediments, carbon dioxide usually serves as the predominant terminal electron acceptor and methanogenic bacteria dominate. Mineralization of PHBO and reference lignocellulose were examined in both marine and freshwater sediments (Figure 3). PHBO was rapidly and extensively mineralized in both sediment types. Mineralization exceeded 90%, and its halflife was less than 6 days in both sediment types (Table 3 and 4). Conversion of PHBO to 14CH4 occurred in both sediment types but at lower levels compared to septage and digester sludge. However, three times more 14CH4 was recovered from freshwater than from marine sediments. Once again, mineralization of lignocelllulose was less extensive than that of PHBO but similar in both sediment types. Total lignocellulose recovered as 14CO2 and 14CH4 from marine and freshwater sediments after 141 days was 35.2% and 36.9%, respectively. More 14CH4 was recovered from freshwater (13.6%) compared to marine (9.6%) sediments. Decomposition of Refuse in Landfill Reactors. A complex series of chemical and biological reactions begins with the burial of MSW in a landfill.43 Initially, aerobic bacteria deplete the oxygen entrained in the refuse, and large amounts of carbon dioxide are produced. Once the oxygen is depleted, there is no mechanism for its replenishment because refuse is typically buried in approximately 3 m thick layers. Thus, anaerobic conditions govern decomposition in landfills. Because there is only a limited supply of alternate electron acceptors such as nitrate and sulfate, methanogenic conditions typically dominate refuse decomposition. Cumulative methane production in reactors containing PHBO and PCL exhibited an exponential increase followed by an asymptote (Figures 4 and 5). This is the classical trend reported in previous work.42 The methane yields from decomposition of the bulk MSW were 89.7 ( 6.6 and 78.2 ( 7.8 mL of CH4 per dry g of fresh refuse (mean ( standard deviation) in the PHBO and PCL reactors, respectively. These yields are well within the range of yields (80.6123.3 mL of CH4/dry g of fresh refuse) measured in studies of this type over the past several years.42,44-47 The pH of the leachates varied initially from 6.1 to 6.4 and increased rapidly to above 6.9 within 18 days of initiation

Figure 3. Mineralization of polyhydroxybutyrate-octanoate (PHBO) and lignocellulose to 14CO2 and 14CH4 in marine and freshwater anaerobic sediments incubated at 22 °C (mean ( standard deviation). The test concentrations were 50 mg/kg. PHBO is represented by the open symbols, and lignocellulose is represented by the closed symbols. Marine sediment is represented by the round symbols and freshwater sediment by the triangular symbols.

of the experiments. The pH was externally neutralized prior to recirculation for the first week, after which neutralization was not necessary. The rapid pH increase can be attributed to the presence of an inoculum of well-decomposed refuse that provided the bacteria required to metabolize the carboxylic acids that are produced as intermediates in anaerobic refuse decomposition. The pH data are also consistent with the observed trend in leachate chemical oxygen demand (COD). Typically, leachate COD starts at relatively high concentrations that result from an accumulation of carboxylic acids. As the activity of the methanogenic, acetogenic, and fermentative bacteria becomes balanced, the concentration of soluble substrate is depleted and the remaining COD is largely recalcitrant organic matter. The reactor data followed this pattern and the COD concentration varied without a discernible trend after day 49, ranging between 500 and 1700 mg/L (data not shown). The combined observations of the methane production rates and yields and the trends in the pH and COD data indicate that the refuse underwent methanogenic decomposition in a pattern consistent with previous studies and that refuse decomposition was consistent across all reactors. Thus, the environmental conditions under which PHBO and PCL



Table 5. Final Distribution (% of added radioactivity) following Incubation of Radiolabeled PHBO for 177 Days and Radiolabeled PCL for 721 Days with Municipal Solid Waste in Simulated Landfill Reactors material reactor PHBO 1 2 3 PCL 1d 2 3

evolved 14CH a 4

evolved 14CO 2

aqueous 14CO 2

14C-organicsa

aqueous

residual refusea

mineralizationa

total recoveryb

11.3 8.2 6.6

40.7 33.1 33.8

0.52 0.65 0.75

1.0 1.2 0.82

19.6 23.0 54.1

52.5 42.0 41.1

73.1 66.2 96.0

14.2 14.7 33.1

19.2 6.7 21.9

7.4 1.8 1.2

0.15 0.11 0.09

c c c

40.8 23.2 56.2

c c c

a Data represent the fraction of the added PHBO recovered as 14CH , 14CO (g), and 14CO (aq). b Data represent the fraction of the total PHBO added 4 2 2 that was recovered in any form. c Not determined. d Reactor destructively sampled after 474 days versus 721 days for other reactors.

Figure 4. Mineralization of polyhydroxybutyrate-octanoate (PHBO) to gaseous 14CH4 and 14CO2 (closed symbols) and total cumulative methane (open symbols) produced in landfill simulation reactors (means ( standard deviations). Dissolved 14CO2 was negligible.

biodegradation was tested were representative of the complete refuse decomposition cycle. Fate of PHBO in Simulated Landfill. The amount of 14 C-PHBO that was converted to 14CH4 or 14CO2 or present in the leachate as dissolved 14CO2 and organic material is summarized in Table 5. A total of 41.1-52.5% of the added PHBO was mineralized to 14CH4 and 14CO2. The majority of the mineralization occurred during the first 30 days of the experiment (Figure 4). PHBO mineralization was more rapid than biodegradation of the bulk MSW as evidenced by the continued increase in methane yield for over 80 days (Figure 4). Methane production is a surrogate for bulk refuse decomposition as methane is largely produced from the decomposition of cellulose and hemicellulose in the refuse.46 From 6 to 12% of the 14C from the added PHBO was present as soluble organic matter in the reactor leachates on day 20, and this amount steadily decreased to less than 2% by day 50 and to between 0.25% and 1% through the duration of the experiment. In contrast, only 0.07% and 0.11% of added 14C was solubilized in an abiotic leaching tests conducted for 99 days with synthetic acid phase and methane phase leachates, respectively. The reactor data suggest that there was some rapid solubilization of PHBO, and the soluble products either accumulated in the reactor leachate or were converted to degradation intermediates. As specific organic analyses were not performed, the composition of the soluble materials is unknown. However, it is quite typical for the

Figure 5. Mineralization of polycaprolactone (PCL) to gaseous 14CH4 and 14CO2 (closed symbols) and total cumulative methane (open symbols) produced in landfill simulation reactors (means ( standard deviations). Dissolved 14CO2 was negligible until roughly day 250 at which point dissolved 14CO2 accounted for an additional 2-7% (see Table 5).

first step in the biodegradation of solid material to involve hydrolysis and solubilization. The amount of radiolabel associated with the solids remaining at the end of the experiment is presented in Table 5. On the basis of these data, the total recovery of 14C was 66-96.0% of the amount added. The most uncertain aspect of the mass balances is the specific activity of the refuse removed from the reactors. Four samples of decomposed refuse from each reactor were analyzed for residual 14C. The coefficient of variation (CV) for these analyses was 25.4, 32.6, and 105.4% for reactors 1, 2, and 3, respectively. The high CVs can be attributed to high sample heterogeneity despite the fact that the entire contents of a reactor were ground to pass a 0.5 mm screen and then mixed prior to combustion analysis for residual 14C. Such heterogeneity is understandable given that 0.06 g of 14C-labeled PHBO was added to approximately 900 g of refuse at the outset and that only 0.1 g of ground refuse could be combusted for analysis of residual radiolabel. Fate of PCL in Simulated Landfill. The amount of 14CPCL that was converted to 14CH4 and 14CO2 as well as the radiolabel present in the leachate as 14CO2 and dissolved organic material is summarized in Table 5. After 200 days, at which time methane production from the bulk refuse was essentially complete, PCL mineralization had reached less than 4% in each reactor (Figure 5). Over the first 70 days, from 0.5 to 1% of the 14C from the added PCL was present

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in the reactor leachate as dissolved organic matter. However, this decreased to less than 0.2% by day 100. This result is consistent with abiotic leaching tests in which only 0.18 and 0.2% of the added PCL was solubilized in synthetic leachate after 90 days at pH 5 and pH 7, respectively. To evaluate whether PCL biodegradation was limited by the presence of an appropriate electron acceptor, sulfate (2 mM) was added to all three reactors on day 220 and to reactors 1 and 3 on days 263 and 290. As illustrated in Figure 5, PCL mineralization began to increase just after day 220. However, as the sulfate was depleted after each addition (data not shown) and no sulfate was added after day 290, it is not clear that sulfate played a role in PCL biodegradation. To further investigate factors controlling PCL biodegradation, reactor 1 was destructively sampled after 474 days, and the refuse was used as an inoculum in batch mineralization assays. Four treatments were tested: (1) fresh PCL, (2) fresh PCL plus 10 mM sulfate, (3) aged PCL, and (4) solubilized PCL. Aged PCL was produced by exposing fresh PCL to synthetic acid phase leachate for 219 days. Solubilized PCL was the 14C released to the aqueous phase during the aging process. Only 1% of the PCL was solubilized during this 219 day leaching period. With the exception of the solubilized PCL treatment, less than 2% of the other treatments were mineralized to 14CO2 and 14CH4 after 235 days. Mineralization of the soluble PCL fraction, however, reached 60.6%. Discussion PHBO Synthesis. Although bacteria are known to produce over 100 different types of PHA, the most common bacterial polyesters generally fall into one of two major types.48 Those with only scl repeat units and those composed of mcl repeat units. More recently, bacteria have been identified that produce copolymers with both scl repeat units and mcl repeat units8,9 such as the PHBO. To incorporate the radiolabel into PHBO, it was expedient to prepare the copolymer by chemical synthesis rather than by fermentation. The use of the new, soluble, dialkylzinc-diol initiator19 facilitated the production of a high molecular weight polymer in the small scale reactions (1-2 g) used for the synthesis of radiolabeled PHBO. Synthetic PHBO is expected to differ only slightly from the equivalent naturally derived PHA. The differences are due to the synthetic copolymer lacking perfect isotacticity. This is a result of the starting 3-alkyl-β-propiolactones being of only 92-95% stereopurity. Similar low levels of imperfection in PHB lead to a 10 °C drop in Tm for synthetic PHB made with the alkylzinc alkoxide initiators compared to the Tm of bacterial PHB. Other differences would be due to differences in molecular weight. However, with Mw of about 100 000, the synthetic PHBO is expected to be representative of natural PHBO for the purposes of biodegradation studies.7 PHBO Biodegradation. The present study shows that PHBO is fully and rapidly biodegraded under anaerobic conditions, and biodegradation should occur in any anaerobic habitat possessing significant microbial activity. The present study differs from previous work reported in the literature

Federle et al.

with regard to the PHA evaluated and the methods utilized. Only one previous report examined the anaerobic biodegradation of a PHA containing mcl monomers and indicated no significant biodegradation of a polymer consisting of 30% 3-hydroxyoctanoate and 70% 3-hydroxydecanoate.25 The current study evaluated the biodegradation of a copolymer containing a mixture of short and medium chain units (90 mol % 3-hydroxybutyrate and 10 mol % hydroxyoctanoate). It is also the first study to utilize radiolabeled polymer for an anaerobic biodegradation study of a PHA allowing direct analysis of 14CO2 and 14CH4 derived from the polymer and the first to systematically evaluate biodegradation under realistic conditions in all relevant anaerobic environments. The anaerobic biodegradability of PHB and copolymers has been demonstrated previously by measuring weight loss or biogas formation compared to a control. This work has been largely limited to polymers containing short-chain monomers derived from biological sources. Several studies have used digester sludge as an inoculum. Budwill et al.20 isolated three types of PHA granules from cultures of Azotobacter Vinelandii grown with different levels of valerate in the medium. PHB and PHBV copolymers containing 13 and 20% 3-hydroxyvalerate were extensively (>80%) converted to methane after 16 days when incubated with an inoculum derived from anaerobically digested domestic sewage sludge. Majone et al.23 incubated PHB granules with a diluted inoculum from an anaerobic digester and measured gas volume and composition as well as volatile fatty acid levels in the medium with time. PHB generated yields of methane comparable to those obtained with glucose, the positive control. In the presence of 2-bromoethanesulfonic acid, an inhibitor of methanogenesis, yields of volatile fatty acids were comparable to those observed with glucose. Puechner et al.24 incubated powdered PHB homopolymer with diluted domestic anaerobic sludge and found that biogas production started after a lag of 3-4 days and reached 80% of that theoretically expected within 30 days. Reischwitz et al.27 examined anaerobic biodegradation of a PHBV copolymer containing 8.4% 3-hydroxyvalerate in a stirred reactor with pH control and gas monitoring containing 10% anaerobic sludge and 1% powdered polymer. After 5 days, 67% of the polymer was converted to organic acids, primarily acetate, and butyrate, and after 30 days, 95% of the carbon was converted to biogas. Other studies using digester sludge have evaluated the anaerobic biodegradation of PHA films rather than granules or powder. Day et al.29 incubated plastic films with an inoculum from an anaerobic reactor initially seeded with sludge from a digester receiving domestic and pulp and paper sludges. PHBV film lost 29% of its mass and 22% of its thickness after 40 days and exhibited a significant increase in methane generation. Mergaert et al.25 incubated injection molded 3 mm thick dogbone-shaped pieces of PHB and PHBV copolymers containing 10% and 20% 3-hydroxyvalerate. After 123 days, the PHB homopolymer exhibited the greatest weight loss (15%) followed by the PHBV copolymer with 10% 3-hydroxyvalerate (11%) and PHBV copolymer with 20% 3-hydroxyvalerate (7%). Shin et al.30 examined the biodegradation of PHBV copolymer films containing 8%



hydroxyvalerate by diluted digester sludge. Within 30 days, 89% of the PHBV was converted to biogas. Gartiser et al.31 incubated PHBV films with diluted digester sludge with different media and purge gases and measured total gas production. Biogas from the PHBV copolymer reached 6075% of theoretical within 30 days. Recently, Abou-Zeid et al.28 examined the biodegradation of films consisting of PHB and PHBV containing 11.6% hydroxyvalerate by sludge from a laboratory anaerobic reactor fed sugar industry wastewater. After 10 weeks, weight loss equaled 23% for the PHB film and 22.5% for the PHBV film. In a second set of experiments PHB exhibited 100% weight loss and 100% of theoretical biogas generation and PHBV exhibited 60% weight loss and 30% biogas formation. Abou-Zeid et al.28 also isolated 26 strains of anaerobic bacteria belonging to the genus Clostridium capable of degrading PHB and PHBV. No previous studies have examined PHA fate in septic tanks or septage, but one examined fate under simulated landfill conditions. Shin et al.30 measured the weight loss of PHBV films containing 8% 3-hydroxyvalerate in a simulated landfill reactor. After 6 months, the PHBV films exhibited 78% weight loss. A few studies have evaluated PHA degradation in sediments or with inocula derived from sediments. Janssen and Hartfoot49 reported the isolation from sediment of an obligate anaerobic bacterium that degrades PHB. Janssen and Schink50 subsequently showed that this anaerobic bacterium, Ilyobacter delafieldii, produced an extracellular depolymerase which released free hydroxybutyrate monomer from PHB that was fermented to acetate, butyrate, and hydrogen. Urmeneta et al.51 observed increased sulfide concentrations and loss of PHBV containing 3% 3-hydroxyvalerate in sediment slurries amended with PHBV. Mas-Castella´ et al.22 incubated known quantities of PHB and PHBV copolymers containing 7% and 14% hydroxyvalerate with anaerobic lake sediments and microbial mats. All three compounds stimulated anaerobic respiration as evidenced by hydrogen sulfide formation and the quantity of hydrogen sulfide formed correlated with the level of PHA added. In addition, little PHA could be recovered from the test systems following incubation. Addition of molybdate, which inhibits sulfate reducing activity, resulted in decreased yields of hydrogen sulfide and higher residual levels of PHA, thereby implicating sulfate-reducing bacteria in PHA biodegradation. Finally, Budwill et al.26 examined anaerobic biodegradation of PHB and PHB copolymemer containing 15% hydroxyvalerate with various terminal electron acceptors and with a variety of inocula. They were able to demonstrate degradation under nitrate-reducing conditions using anoxic-activated sludge cultures based upon stimulated CO2 production and nitrate consumption. Under methanogenic conditions, addition of PHA to cultures derived from pond sediment stimulated methane production, but no stimulation in methane production was observed with rumen fluid. No evidence of PHA degradation was observed with iron or sulfate reducing cultures derived from springwater. PCL Biodegradation. In the present study, PCL was not significantly biodegraded in digester sludge after 122 days or in the landfill simulation after 200 days. Once again biodegradation was based upon specific measurement of

14

CO2 and 14CH4 derived from the radiolabeled polymer incubated under realistic conditions with actual environmental samples. Previous studies also have observed little or minimal biodegradation of PCL. Day et al.29 incubated plastic films with an inoculum from an anaerobic reactor initially seeded with sludge from a digester receiving domestic and pulp and paper sludges. PCL film exhibited only a 6% loss in mass and no significant loss in thickness or increase in methane production. Gartiser et al.31 incubated PCL films with diluted digester sludge with different media and purge gases and measured total gas production. No significant production of biogas was observed with PCL after 80 days. Abou-Zeid et al.28 examined the biodegradation of PCL film by sludge from a laboratory anaerobic reactor fed sugar industry wastewater. After 10 weeks weight loss equaled 7.6%, and in a second experiment PCL showed 30% weight loss and 17% biogas production. Abou-Zeid et al.28 also isolated two bacterial strains related to Clostridium acetobutylicum capable of degrading PCL. One of the isolates degraded PCL at a maximum rate of 0.14 mg/week in the presence of 40-60 mg of PCL, representing a rate of only 0.2-0.3% per week. In the landfill study, significant mineralization of PCL began only after 200 days and continued over the next 500 days. There are three potential explanations for this late onset of PCL biodegradation. First, PCL undergoes slow hydrolysis yielding biodegradable products. Second, a long acclimation period was required for the microbial community to acquire the ability to degrade PCL. Third, the microbial populations responsible for PCL and MSW biodegradation were the same and that MSW was used preferentially. The last two explanations are unlikely given the results of the serum bottle assays. Little mineralization was observed in these tests, which were conducted with bacteria acclimated to PCL and with PCL as the sole carbon source. Given the results with solubilized PCL, it appears that hydrolyzed PCL is biodegradable, and hydrolysis, whether abiotic or biologically mediated is very slow and limits the anaerobic biodegradation of PCL. Inherently biodegradable means that a material can be completely biodegraded under some circumstances. A material that is practically biodegradable is completely biodegraded in the habitats to which it is discarded and its inherent biodegradability is realized. Furthermore, the biodegradation rate of a practically biodegradable polymer exceeds or is similar to the loading rate of the material into the environment, such that biodegradation is a meaningful removal mechanism and the polymer does not accumulate. Upon the basis of testing in digester sludge, septage, sediments, and landfills, PHBO is practically biodegradable in the major anaerobic habitats in which it could be discarded. In contrast, PCL does not meet the definition of practically biodegradable in anaerobic sewage sludge given typical digester residence times. PCL, however, may be practically biodegradable under anaerobic landfill conditions given the long residence time and results from the laboratory landfill simulation. Nevertheless, this conclusion remains equivocal. It is not known to what degree radiolysis may have contributed to the hydrolysis of the radiolabeled PCL used in this study given its long

822


duration. It is possible that the presence of radiolabeled atoms within the PCL test material could destabilize the molecule and contribute to its hydrolysis. Thus, further confirmation will be required before it can be concluded that PCL is practically biodegradable in landfill environments. In this study, the polymers were tested in a powdered form. However, the physical form that a biodegradable plastic takes can significantly influence the kinetics and possibly the occurrence of biodegradation. Although a product may be made from a practically biodegradable material, there is no guarantee that the product itself will be practically biodegradable. Consequently, in assessing the anaerobic biodegradability of product, it is important to verify that the product itself undergoes disintegration in the anaerobic environment in a reasonable time. These data combined with data such as that generated in this study can then be combined to definitively establish the practical biodegradability of a particular product or article. Even for articles made with PHBO, which is unequivocally anaerobically biodegradable, a need will exist to assess disintegration of the articles themselves. This process will be particularly challenging for objects made from PCL, which has a less definitive anaerobic biodegradation profile. Acknowledgment. The authors thank Messrs. G. M. Bunke and J. Innis for performing the radiolabel syntheses. References and Notes (1) Characterization of Municipal Solid Waste in the United States: 1998 Update; Office of Solid Waste, U.S. Environmental Protection Agency: Washington, DC, 1999. (2) Bastioli, C. Macromol. Symp. 1998, 135, 193. (3) Muller, H.-M.; Seebach, D. Angew. Chem., Int. Ed. Engl. 1993, 32, 477. (4) Williams, S. F.; Peoples, O. P. CHEMTECH 1996, 26 (9), 38. (5) Doi, Y. Microbial Polyesters; VCH Publishers: New York, 1990. (6) Doi, Y.; Kitamura, S.; Abe, H. Macromolecules 1995, 28, 4822. (7) Satkowski, M. M.; Melik, D. H.; Autran, J.-P.; Green, P. R.; Noda, I.; Schechtman, L. A. In Biopolymers; Steinbu¨chel, A., Doi, Y., Eds.; Wiley: New York, 2002; Volume 3B, Polyesters I, Chapter 9. (8) Matsusaki, H.; Abe, H.; Doi, Y. Biomacromolecules 2000, 1, 17. (9) Green, P. R.; Kemper, J. J.; Schechtman, L. A.; Guo, L.; Satkowski, M. M.; Fiedler, S.; Steinbuechel, A.; Rehm, B. H. A. Biomacromolecules 2002, 3, 208-213. (10) Benvenuti, M.; Lenz, R. W. J. Polym. Sci., Part A: Polym. Chem. 1991, 29, 793. (11) Isoda, M.; Sugimoto, H.; Aida, T.; Inoue, S. Macromolecules 1997, 30, 57. (12) Lenz, R. W.; Jedlinski, Z. Macromol. Symp. 1996, 107, 149. (13) (a) Kricheldorf, H. R.; Lee, S.-R.; Scharnagl, N. Macromolecules 1994, 27, 3139. (b) Kricheldorf, H. R.; Lee, Macromolecules 1995, 28, 6718. (c) Kemnitzer, J. E.; McCarthy, S. P.; Gross, R. A. Macromolecules 1993, 26, 6143. (d) Kemnitzer, J. E.; McCarthy, S. P.; Gross, R. A. Macromolecules 1993, 26, 1221. (14) Hori, Y.; Suzuki, M.; Yamaguchi, A.; Nishishita, T. Macromolecules 1993, 26, 5533. (15) Le Borgne, A.; Pluta, C.; Spassky, N. Macromol. Rapid Commun. 1994, 15, 955. (16) Zhang, Y.; Gross, R. A.; Lenz, R. W. Macromolecules 1990, 23, 3206.

Federle et al. (17) Le Borgne, A.; Spassky, N. Polymer 1989, 30, 2312. (18) (a) Schechtman, L. A.; Kemper, J. J. WO 95/20616, 1995; Chem. Abstr. 1995, 124, 9835. (b) Schechtman, L. A.; Kemper, J. J. U.S. Patent 5648452, 1997. (c) Schechtman, L. A.; Kemper, J. J. Polym. Prepr. 1999, 40 (1), 508. (19) Schechtman, L A.; Kemper, J. J. WO 00/77072, 2000; Chem. Abstr. 2000, 134, 57105. (20) Budwilll, K.; Fedorak, P. M.; Page, W. J. Appl. EnViron. Microbiol. 1992, 58, 1398. (21) Janssen, P. H.; Schink, B. Biodegradation 1993, 4, 179. (22) Mas-Castella`, J.; Urmenata, J.; Lafuente, R.; Navarrete, A.; Guerrero, R. Int. Biodeterior. Biodegrad. 1995, 35, 155. (23) Majone, M.; Riccardi, C.; Rolle, E.; Scarinci, A. Toxicol. EnViron. Chem. 1995, 48, 103. (24) Puechner, P.; Mueller, W.-R.; Bardtke, D. J. EnViron. Polym. Degrad. 1995, 3, 133. (25) Mergaert, J.; Glorieux, G.; Huben, L.; Storms, V.; Mau, M.; Swings, J. System. Appl. Microbiol. 1996, 19, 407. (26) Budwill, K.; Fedorak, P. M.; Page, W. J. J. EnViron. Polym. Degrad. 1996, 4, 91. (27) Reischwitz, A.; Stopok, E.; Buchholz, K. Biodegradation 1998, 8, 313. (28) Abou-Zeid, D.-M.; Mu¨ller, R.-J.; Deckwer, W.-D. J. Biotechnol. 2001, 86, 113. (29) Day, M.; Shaw, K.; Cooney, D. J. EnViron. Polym. Degrad. 1994, 2, 121. (30) Shin, P. K.; Kim, M. H.; Kim, J. M. J. EnViron. Polym. Degrad. 1997, 5, 33. (31) Gartiser, S.; Wallrabenstein, M.; Stiene, G. J. EnViron. Polym. Degrad. 1998, 3, 159. (32) Goldberg, D. J. EnViron. Polym. Degrad. 1995, 3 (2), 61-67. (33) (a) Benedict, C. V.; Cameron, J. A.; Huang, S. J. J. Appl. Polym. Sci. 1983, 28, 335. (b) Benedict, C. V.; Cook, W. J.; Jarrett, P.; Cameron, J. A.; Huang, S. J.; Bell, J. P. J. Appl. Polym. Sci. 1983, 28, 327. (34) Pettigrew, C. A.; Rece, G. A.; Smith, M. C.; King, L. W. J. Macrobiol. Sci. Pure Appl. Chem. 1995, A32, 811. (35) Ohta, T.; Miyake, T.; Takaya, H. J. Chem. Soc., Chem. Commun. 1992, 1725. (36) (a) Capozzi, G.; Roelens, S.; Talami, S. J. Org. Chem. 1993, 58, 7932. (b) Breitshuh, R.; Seebach, D. Chimia 1990, 44, 216. (c) Sato, T.; Kawara, T.; Nishizawa, A.; Fujisawa, T. Tetrahedron Lett. 1980, 21, 3377. (37) Crawford, D. L.; Crawford, R. L. Appl. EnViron. Microbiol. 1976, 31, 714. (38) Crawford, R. L. Lignin Biodegradation and Transformation; John Wiley and Sons: New York, 1981; pp 20-37. (39) Effland, M. J. Tappi 1977, 60, 143. (40) Jacobs, C.; Dubois, P.; Jerome, R.; Teyssie, P. Macromolecules 1991, 24, 3027. (41) Nuck, B. A.; Federle, T. W. EnViron. Sci. Technol. 1996, 30, 3597. (42) Ress, B. B.; Calvert, P. P.; Pettigrew, C. A.; Barlaz, M. A. EnViron. Sci. Technol. 1998, 32, 821. (43) Barlaz, M. A.; Schaefer, D. M.; Ham, R. K. Appl. EnViron. Microbiol. 1989, 55, 55. (44) Eleazer, W. E.; Odle, W. S.; Wang, Y.-S.; Barlaz, M. A. EnViron. Sci. Technol. 1997, 31, 911. (45) Fairweather, R. J.; Barlaz, M. A. J. EnViron. Eng. 1998, 124, 353. (46) Barlaz, M. A.; Ham, R. K.; Schaefer, D. M. J. EnViron. Eng. 1989, 115, 1088. (47) Parkin, G. F.; Owen, W. F. J. EnViron. Eng. 1986, 112, 867. (48) Steinbu¨chel A. In Biotechnology; Rehm, H. J., Reed, G., Eds.; VCH: Weinheim, 1996; pp 403-464. (49) Janssen, P. H.; Hartfoot, C. G. Arch. Microbiol. 1990, 154, 253. (50) Janssen, P. H.; Schink, B. Biodegradation 1993, 4, 179. (51) Urmenata, J.; Mas-Castell, J.; J.; Guerrero, R. Appl. EnViron. Microbiol. 1995, 61, 2046.

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Anaerobic Biodegradation of Aliphatic Polyesters - American

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