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TECHNICAL REPORTS

Bioremediation and Biodegradation

Biodegradation of Polycyclic Aromatic Hydrocarbons in Oil-Contaminated Beach Sediments Treated with Nutrient Amendments

Department of Chemical and Environmental Engineering, National University of Singapore, 10 Kent Ridge Crescent, Singapore 119260

^* Corresponding author (chejpo{at}nus.edu.sg).

Received for publication June 20, 2003.

	ABSTRACT

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Microbial biodegradation of polycyclic aromatic hydrocarbons (PAHs) during the process of bioremediation can be constrained by lack of nutrients, low bioavailability of the contaminants, or scarcity of PAH-biodegrading microorganisms. This study focused on addressing the limitation of nutrient availability for PAH biodegradation in oil-contaminated beach sediments. In our previous study, three nutrient sources including inorganic soluble nutrients, the slow-release fertilizer Osmocote (Os; Scotts, Marysville, OH) and Inipol EAP-22 (Ip; ATOFINA Chemicals, Philadelphia, PA), as well as their combinations, were applied to beach sediments contaminated with an Arabian light crude oil. Osmocote was the most effective nutrient source for aliphatic biodegradation. This study presents data on PAH biodegradation in the oil-spiked beach sediments amended with the three nutrients. Biodegradation of total target PAHs (two- to six-ring) in all treatments followed a first-order biodegradation model. The biodegradation rates of total target PAHs in the sediments treated with Os were significantly higher than those without. On Day 45, approximately 9.3% of total target PAHs remained in the sediments amended with Os alone, significantly lower than the 54.2 to 58.0% remaining in sediment treatments without Os. Amendment with Inipol or soluble nutrients alone, or in combination, did not stimulate biodegradation rates of PAHs with a ring number higher than 2. The slow-release fertilizer (Os) is therefore recommended as an effective nutrient amendment for intrinsic biodegradation of PAHs in oil-contaminated beach sediments.

Abbreviations: Ip, Inipol EAP-22 • Os, Osmocote • PAH, polycyclic aromatic hydrocarbon • SN, soluble nutrients

	INTRODUCTION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
OIL AND OIL-CONTAMINATED environmental matrices are extremely complex mixtures that contain a vast array of aliphatic and polycyclic aromatic hydrocarbons (PAHs) (Wang et al., 1994). Polycyclic aromatic hydrocarbons are toxic and hazardous chemicals regulated by the USEPA as priority pollutants (Cho and Kim, 1997; Tabak et al., 1997; Juhasz and Naidu, 2000). Therefore, it is important to remove PAHs from the environment both quickly and safely following an oil spill incident. Bioremediation can be an effective option to reclaim PAH-contaminated sites due to its relatively low cost and limited impact on the environment (Liebeg and Cutright, 1999).

Polycyclic aromatic hydrocarbons have low water solubilities and tend to bind with organic matter or particle surfaces, resulting in a low bioavailability to the microbial biomass (Cerniglia, 1992; Taylor and Jones, 2001). In addition, PAHs are thermodynamically stable since they are derivatives of the benzene ring with large negative resonance energies (Mueller et al., 1996). As a result, PAHs are recalcitrant in the environment and are often resistant to biodegradation under prevailing natural conditions. High-ring-number PAHs are more difficult to biodegrade than one- and two-ring aromatics, where condensed-ring aromatic hydrocarbons are highly resistant to enzymatic attack (Cerniglia, 1992; Mueller et al., 1996). Acceleration of the PAH biodegradation process can be achieved by manipulating the substrate microenvironment, such as by adding nutrients (Oh et al., 2001), enhancing aerobic status (Symons et al., 1995), introducing microbial inoculum (Rahman et al., 2002; Tam et al., 2002), or enhancing PAH bioavailability (Barkay et al., 1999; Duke et al., 2000; Bogan et al., 2003).

This study is a supplement to our previous work that investigated the effect of bioremediation additives on the biodegradation of aliphatics in oil-contaminated beach sediments under tropical marine conditions (Xu and Obbard, 2003). In the previous study, the addition of the slow-release inorganic fertilizer, Osmocote (Os), to sediment stimulated the indigenous biodegradation of aliphatics more effectively than soluble nutrients (SN) or Inipol EAP-22 (Ip). In this manuscript we report the ability of these nutrient sources to enhance PAH biodegradation in oil-contaminated beach sediments. Specifically, we investigated the biodegradation of two- to six-ring PAHs, as well as the C1 to C4 alkyl homologs of two- and three-ring PAHs, by the indigenous microbial biomass in an oil-spiked beach sediment.

	MATERIALS AND METHODS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
Experimental Setup
The experimental setup, materials, and methods used for biological and nutrient analyses, as well as analyses for aliphatic petroleum hydrocarbons [i.e., n-C₁₂ to n-C₃₃, pristane, phytane, and C₃₀–17(H), 21ß(H)-hopane], have been reported in detail in our previous study (Xu and Obbard, 2003). The uncontaminated beach sediment used for this study was collected from Pulau Semakau, a small island 8 km south of Singapore and subsequently spiked in the laboratory with an Arabian light crude oil to 3.22% (w/w, dry weight equivalent). After three weeks of weathering, the oil content decreased to 2.20%. The experimental setup was designed as an "open" irrigation system, where sediment was free-draining and irrigated twice per day. Three forms of nutrient amendment, including SN (a mixture of NH₄NO₃, K₂HPO₄, and KH₂PO₄), Inipol EAP-22 (26.2% oleic acid, 23.7% lauryl phosphate, 10.8% 2-butoxy-1-ethanol, 15.7% urea, and 23.6% water), and Osmocote 18–11–10 (subsequently referred to as Os) were amended to the oil-spiked sediment. Osmocote contains water-soluble N–P–K at concentrations of 18, 4.8, and 8.3% (w/w) respectively, with no other trace constituents. The control comprised a biotic, non-amended, oil-spiked sediment.

A range of treatments (Table 1) were performed in duplicate. Sediment sampling (25 g per microcosm) for hydrocarbon analyses was undertaken on Days 0, 7, 15, 30, and 45, and for biological analysis on Days 0, 2, 6, 12, 21, 30, 38, and 45. Seawater leachate from the sediments was collected and analyzed for nutrient levels on Days 1, 2, 7, 15, 25, and 45.

Each sediment sample (5 g, dry weight equivalent) was extracted by 165 mL hexane–acetone (1:1, v/v) mixture using a Soxhlet extraction system (Eaton et al., 1995, p. 5-34 to 5-35). The extract was filtered and analyzed on a gas chromatograph–mass spectrometer (GC–MS) for alkanes (n-C₁₂ to n-C₃₃, pristane, and phytane), C₃₀–17 (H), 21ß(H)-hopane (Xu and Obbard, 2003), and target PAHs (two- to six-ring PAHs, as well as the C1 to C4 alkyl homologs of two- and three-ring PAHs). The Hewlett-Packard (Palo Alto, CA) 6890 GC used in this study was equipped with a HP 6890 series mass selective detector (Model 5972A), a HP 6890 autosampler, and a HP 19091S-433, HP-5MS 5% phenyl methyl siloxane, 30.0-m x 250-µm i.d. (0.25-µm film) capillary column. The flow rate of carrier gas, helium, was at 1.6 mL min^–1. The injector and detector temperatures were set at 290 and 300°C, respectively.

The temperature program for target PAHs was: 1-min hold at 90°C, ramp to 160°C at 25°C min^–1, ramp to 290°C at 8°C min^–1, and 15-min hold at 290°C. The temperature program for the alkyl homologs of PAHs was: 2-min hold at 50°C, ramp to 300°C at 6°C min^–1, and 16-min hold at 300°C. A 1-µL aliquot of solvent was injected into the GC–MS using a splitless mode with a 6-min purge-off. The mass selective detector was operated in the scan mode to obtain spectral data for identification of hydrocarbon components and in the selected ion monitoring (SIM) mode for quantification of target compounds. Ions monitored for PAH analysis are summarized in Table 2.

[1]

where A is the concentration of analyte, H is the concentration of hopane, k is the first-order biodegradation rate constant for the analyte, (A/H) is the time-varying hopane normalized concentration of the analyte, and (A/H)₀ is the theoretical value of that quantity at the onset of biodegradation. Normalizing Eq. [1] by ^*₀, which is the experimental value of (A/H) at time 0, yields the following first-order relationship:

[2]

Simplifying Eq. [2] gives the following relationship:

[3]

where:

[4]

Nonlinear regression analysis was used to estimate the first-order rates (k), the coefficients of determination (r²), and the y intercepts of PAH biodegradation for each of the seven sediment treatments to determine how well the experimental data approximated to the first-order.

Tukey's one-way analysis of variance (ANOVA) test at a family error rate of 5% was used to determine the statistical significance of chemical and biological data. Data were considered to be significantly different between two values if p < 0.05. All statistical analyses were performed using Minitab Release 13.20 (Minitab, 2000).

	RESULTS AND DISCUSSION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
In our previous study, nitrogen and phosphorus concentrations in seawater leachate from SN, Ip, and Ip + SN sediments decreased significantly in the initial 15 d and then remained at a relatively stable and low level for the remaining period of the 45-d experiment. In contrast, the nutrient concentrations in leachate from the sediments treated with Os (i.e., Os, SN + Os, and Ip + Os) increased significantly in the first 15 d and subsequently remained at significantly higher levels than the other treatments and the control (p < 0.05). The metabolic activity of the microbial biomass, as measured by intracellular dehydrogenase activity in the sediments either amended or unamended with nutrients, corresponded with the variation of nutrient concentration in the leachate. As a result, the dehydrogenase activity was significantly higher in sediments treated with Os than those without (p < 0.05).

In this study, we found that the biodegradation of total target PAHs (i.e., two- to six-ring PAHs and the C1 to C4 alkyl homologs of two- and three-ring PAHs) and total target alkanes (n-C₁₂ to n-C₃₃, pristane and phytane) followed the first-order decline model (Eq. [3]). The derived values for the first-order rate constant (k), coefficient of determination (r²), normalized y intercept of total target PAH, alkane, and two-ring PAH biodegradation are summarized in Table 3. A summary of alkane data is presented so that a relative comparison can be made between alkane and PAH losses from sediments. Biodegradation of alkanes in oil-contaminated sediments was reported in our previous manuscript (Xu and Obbard, 2003). Figure 1 shows the biodegradation of total target PAHs in the control and nutrient-treated sediments. Figures 2 through 6 show the biodegradation of individual PAHs with ring numbers from 2 to 6. The approximated first-order biodegradation kinetics based on the total PAH loss data are also shown in Fig. 1.

View larger version (25K): [in this window] [in a new window]   Fig. 1. Approximation of actual polycyclic aromatic hydrocarbon (PAH) loss (symbols) and first-order loss kinetics (lines) for total target PAHs (i.e., two- to six-ring PAHs and C1 to C4 alkyl homologs of two- and three-ring PAHs).

View larger version (27K): [in this window] [in a new window]   Fig. 2. Degradation of total target two-ring polycyclic aromatic hydrocarbons (PAHs) (i.e., naphthalene and its C1 to C4 alkyl homologs) relative to hopane for the different treatments over time.

View larger version (23K): [in this window] [in a new window]   Fig. 6. Degradation of total target six-ring polycyclic aromatic hydrocarbons (PAHs) relative to hopane for the different treatments over time.

View larger version (26K): [in this window] [in a new window]   Fig. 3. Degradation of total target three-ring polycyclic aromatic hydrocarbons (PAHs) and their C1 to C4 alkyl homologs relative to hopane for the different treatments over time.

View larger version (23K): [in this window] [in a new window]   Fig. 4. Degradation of total target four-ring polycyclic aromatic hydrocarbons (PAHs) relative to hopane for the different treatments over time.

View larger version (24K): [in this window] [in a new window]   Fig. 5. Degradation of total target five-ring polycyclic aromatic hydrocarbons (PAHs) relative to hopane for the different treatments over time.

The biodegradation rates of total target PAHs in all Os-treated samples were significantly higher than those without (p < 0.05; Table 3), where the mean loss rate of the former was approximately 2.9-fold higher than the latter as a result of biodegradation. However, there was no significant difference in the biodegradation rates of total target PAHs between sediments treated with Ip, SN, and Ip + SN, as well as the oil-spiked control (p > 0.05). Neither was there a significant difference between the treatments with Os alone and those with Os combined with Ip or SN (p > 0.05). At the end of the experiment, on Day 45, only 9.3, 9.0, and 19.0% of total target PAHs remained in the sediments treated with Os alone, SN + Os, and Ip + Os, respectively (see Fig. 1). In contrast, 54.2 to 58.0% of total target PAHs was still present in the oil-spiked control sediment and those treated with Ip alone, SN alone, or in combination (Ip + SN) (see Fig. 1). Therefore, addition of Ip and SN to oil-spiked sediments had no additional benefit on the biodegradation of total target PAHs over that of Os alone in this 45-d experiment. In contrast, the presence of SN plus Os was found to be favorable for the stimulation of less recalcitrant aliphatics, as reported previously (Xu and Obbard, 2003).

Biodegradation of Two-Ring Polycyclic Aromatic Hydrocarbons
From Table 3, it can be seen that the r² values for two-ring PAH (i.e., naphthalene and its C1 to C4 alkyl homologs) losses in sediments treated with Os are higher than 0.95, meaning that the biodegradation of two-ring PAHs in these sediments complied with first-order reaction kinetics. The first-order biodegradation rates of two-ring PAHs for all treatments with Os were high and similar to corresponding alkane biodegradation rates (see Table 3). This may be due to their higher bioavailability to the microbial biomass compared with PAHs of a higher ring number. It has previously been reported that some simple aromatics (i.e., naphthalene and 2-methylnaphthalene) can be biodegraded even more rapidly than alkanes (Fedorak and Westlake, 1981a, 1981b). From Fig. 2, it can be seen that the concentration of two-ring PAHs reached a plateau on Day 15 in SN and Ip + SN sediments, and on Day 30 in Ip sediments. This is associated with the significant decrease in nutrient concentration due to leaching in these sediments over the initial 15 d of the study (see Xu and Obbard, 2003). As a result, the biodegradation of two-ring PAHs in the above treated sediment does not follow first-order kinetics.

In our study, 48% of naphthalene and its C1 to C4 alkyl homologs still remained in the oil-spiked control sediments after 45 d, and concentrations were significantly higher (p < 0.05) than all treated sediments (2.8–41%). Among the treated sediments, 41% of the initial two-ring PAHs remained in the sediments treated with Ip alone, and this was significantly higher (p < 0.05) than all other treatments (2.8–27.4%). In addition, the biodegradation rate of two-ring PAHs in the Ip + Os sediment was significantly lower than Os and SN + Os sediments. On Day 45, 25.4% of the two-ring PAHs remained in the Ip + Os sediment, significantly higher than the 3.8 and 2.8% in Os and SN + Os sediments, respectively (Fig. 2; p < 0.05). As for total alkane biodegradation, this may be due to the presence of oleic acid in Ip acting as an alternative carbon source.

Biodegradation of Three- to Six-Ring Polycyclic Aromatic Hydrocarbons
The low solubility and volatility of many higher molecular weight PAHs makes them less susceptible to physical loss processes. As a result, their loss may be more readily attributed to biodegradation processes (Pritchard et al., 1992). Typically, high-ring-number PAHs are more difficult to biodegrade than one- and two-ring aromatics, and condensed ring aromatic hydrocarbons are even more resistant to enzymatic attack (Atlas and Bartha, 1992). Figures 3 through 6 show the biodegradation of three- to six-ring PAHs in the oil-contaminated sediments treated with the various fertilizers, and the results are consistent with this trend.

The r² values for three- to six-ring PAHs in the sediments treated with Os are generally below 0.90 and do not approximate to first-order reaction kinetics especially for four-ring PAHs. However, from Fig. 3 through 6, it can be seen that (i) the biodegradation rate of PAHs declined with increasing benzene ring number in all treatments, (ii) biodegradation rates of PAHs with ring numbers of three to six in sediments treated with Os were obviously higher than all other treatments without Os, and (iii) there were no obvious differences in PAH biodegradation rates for all treatments amended with Os, neither for those without Os. From Fig. 3 through 5, it can be seen that the biodegradation rates of target PAHs in Ip + Os treated sediments were more rapid than in all other sediments in the initial 30 d, before complete Ip leaching loss from sediments. This is probably due to the surfactant properties of Ip enhancing the bioavailability of three- to five-ring PAHs. In contrast, Ip had no obvious effect on the biodegradation of six-ring PAHs, which are more recalcitrant. At the end of the experiment, on Day 45, the percentages of three-, four-, five-, and six-ring PAHs remaining in the sediments treated with Os were 8.7 to 9.0% (Fig. 3), 14.1 to 15.0% (Fig. 4), 42.8 to 45.6% (Fig. 5), and 55.7 to 64.5% (Fig. 6), respectively. These levels were significantly lower compared sediments without Os (p < 0.05), including the oil-spiked control, at 57.9 to 67.3% for three-ring PAHs (Fig. 3), 93.2 to 98.8% for four-ring PAHs (Fig. 4), 74.4 to 91.0% for five-ring PAHs (Fig. 5), and 85.6 to 91.8% for six-ring PAHs (Fig. 6)

The deviation of PAH (i.e., three- to six-ring) biodegradation in sediments treated with Os from the first-order kinetics may be explained by cometabolism. Cometabolism is a term used to describe the process in which a microorganism utilizes a readily degradable substrate as the carbon (energy) source to degrade an organic compound that it is unable to use as a sole carbon (energy) source (Zhang et al., 1998). To date, no research on oil biodegradation has shown that high molecular weight PAHs can be readily utilized as a carbon or energy source by the indigenous microbial biomass. The disappearance of a low concentration of PAHs is likely to be first order. Other hydrocarbons are present in oil-contaminated sediments that support the growth of microorganisms. Microbial metabolism of PAHs is a function of both the growth kinetics of these secondary organic compounds and the kinetics that apply to enzyme systems catalyzing the metabolism of PAHs. Hence, the loss of PAHs will not necessarily follow that of first-order kinetics. For example, the breakdown of low concentrations of phenol or glucose by two bacteria growing on other C sources is best fit by a logistic and a logarithmic loss model (Alexander, 1999, p. 94–97).

Venosa et al. (1996) undertook a field study on bioremediation of an experimental oil spill on the shoreline of Delaware Bay, USA. The SN, NaNO₃, and Na₅P₃O₁₀ were applied intensively to the sediments on a daily basis at above the optimal concentration to counteract nutrient loss via sediment leaching. A first-order loss rate of 0.031 d^–1 was achieved for total target PAHs, which is similar to the result of 0.033 d^–1 in our study. However, as Osmocote is a slow-release fertilizer and was applied only once at 1.9% (sediment dry-weight equivalent), at the beginning of our experiment, a high level of biodegradation was sustained. In our previous report for this experiment (Xu and Obbard, 2003), it was shown that a single application of Osmocote was able to maintain an elevated level of soluble nutrients in a leached sediment over the entire duration of the experiment. This has operational advantages for a bioremediation program of oil-contaminated beach sediments in the field.

The slower biodegradation of high-ring-number PAHs (i.e., four- to six-ring PAHs) in sediments without Osmocote may be explained by the lower prevailing concentration of nutrients in these sediment treatments (Xu and Obbard, 2003). Polycyclic aromatic hydrocarbons in crude oil are invariably present with other aliphatic hydrocarbons together with low-ring-number PAHs, which are more readily biodegradable. Biodegradation of these co-contaminants can be expected to create a biological oxygen and nutrient demand, thereby curtailing biodegradation of the more recalcitrant higher-ring-number PAHs (Mueller et al., 1996).

	CONCLUSIONS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 
The presence of Osmocote at a concentration of 1.9% (sediment dry-weight equivalent) was able to sustain PAH biodegradation in the leached oil-contaminated beach sediment. The biodegradation rates of total target PAHs (i.e., two- to six-ring PAHs and the C1 to C4 alkyl homologs of two- and three-ring PAHs) were significantly higher in all sediments treated with Osmocote than in an oil-spiked control (p < 0.05). In all treated sediments without Os, the presence of Ip or SN alone, or in combination, did not stimulate the biodegradation rates of target PAHs significantly (p > 0.05).

Choosing a suitable form of nutrient amendment can significantly enhance indigenous microbial biodegradation of oil components, including alkanes and PAHs in leached beach sediments. Our study advocates the use of slow-release fertilizers, such as Osmocote, as a sustained and effective source of nutrients for the intrinsic bioremediation of oil-contaminated beach sediments.

	REFERENCES

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS AND DISCUSSION CONCLUSIONS REFERENCES

 

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