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Applied and Environmental Microbiology, July 2001, p. 3127-3133, Vol. 67, No. 7
Institute of Microbiology, University of
Innsbruck, A-6020 Innsbruck, Austria
Received 18 December 2000/Accepted 24 April 2001
We investigated the feasibility of bioremediation as a treatment
option for a chronically diesel-oil-polluted soil in an alpine glacier
area at an altitude of 2,875 m above sea level. To examine the
efficiencies of natural attenuation and biostimulation, we used
field-incubated lysimeters (mesocosms) with unfertilized and fertilized
(N-P-K) soil. For three summer seasons (July 1997 to September 1999),
we monitored changes in hydrocarbon concentrations in soil and soil
leachate and the accompanying changes in soil microbial counts and
activity. A significant reduction in the diesel oil level could be
achieved. At the end of the third summer season (after 780 days), the
initial level of contamination (2,612 ± 70 µg of hydrocarbons g
[dry weight] of soil Bioremediation of
hydrocarbon-contaminated soils, which exploits the ability of
microorganisms to degrade and/or detoxify organic contamination, has
been established as an efficient, economic, versatile, and
environmentally sound treatment (27). On-site-off-site and in situ systems may be used. Decontamination of polluted sites in
cold climates has received increasing interest recently. Considerable oil bioremediation potential has been reported for a variety of terrestrial and marine cold ecosystems, including arctic, alpine, and
antarctic soils; Alaskan groundwater; and antarctic seawater and sea
ice (reviewed in references 3 and 19). Field temperatures play a significant role in controlling the nature and extent of hydrocarbon metabolism. Temperature affects the rate of biodegradation, as well as the physical nature and chemical composition of hydrocarbons (4, 36).
Monitored natural attenuation (intrinsic bioremediation) is becoming
the accepted option for low-risk oil-contaminated sites and is a
cost-effective remediation alternative (13) as it has few
costs other than monitoring costs and the time required for natural
processes to proceed (27). Biodegradation is most often the primary mechanism for contaminant destruction; however, physical and chemical processes, such as dispersion, dilution, sorption, volatilization, and abiotic transformations, are also important (33). The most widely used bioremediation procedure is
biostimulation of the indigenous microorganisms by addition of
nutrients, as input of large quantities of carbon sources (i.e.,
contamination) tends to result in rapid depletion of the available
pools of major inorganic nutrients, such as N and P (26).
Several studies of the effects of biostimulation with mainly N-P-K or
oleophilic fertilizers have reported positive effects on oil
decontamination in cold ecosystems (reviewed in references 3 and
19).
The objective of our study was to determine the feasibility of
bioremediation as a treatment option for a chronically
diesel-oil-polluted soil in an alpine glacier area at an altitude of
2,875 m above sea level. Oil pollution in ski resorts is caused by the
use of motor vehicles for preparation of ski runs and also by leaks and storage tank ruptures. To examine the efficiencies of natural attenuation and biostimulation, we used field-incubated lysimeters (mesocosms) with unfertilized and fertilized soil. For three summer seasons (July 1997 to September 1999), we monitored changes in hydrocarbon concentrations in soil and soil leachate and the
accompanying changes in soil microbial counts and activity.
Study site.
The field study was performed on the
Eisgratferner Glacier in the Tyrolean Stubai Alps at an altitude of
2,875 m above sea level. The mean annual air temperatures in this area
were 0.6,
0099-2240/01/$04.00+0 DOI: 10.1128/AEM.67.7.3127-3133.2001
Copyright © 2001, American Society for Microbiology. All rights reserved.
Bioremediation (Natural Attenuation and
Biostimulation) of Diesel-Oil-Contaminated Soil in an Alpine Glacier
Skiing Area
ABSTRACT
Top
Abstract
Introduction
Materials and Methods
Results and Discussion
References
1) was reduced by (50 ± 4)% and
(70 ± 2)% in the unfertilized and fertilized soil, respectively.
Nonetheless, the residual levels of contamination (1,296 ± 110 and 774 ± 52 µg of hydrocarbons g [dry weight] of
soil1 in the unfertilized and fertilized soil,
respectively) were still high. Most of the hydrocarbon loss occurred
during the first summer season ([42 ± 6]% loss) in the fertilized
soil and during the second summer season ([41 ± 4]% loss) in the
unfertilized soil. In the fertilized soil, all biological parameters
(microbial numbers, soil respiration, catalase and lipase activities)
were significantly enhanced and correlated significantly with each
other, as well as with the residual hydrocarbon concentration, pointing
to the importance of biodegradation. The effect of biostimulation of the indigenous soil microorganisms declined with time. The microbial activities in the unfertilized soil fluctuated around background levels
during the whole study.
INTRODUCTION
Top
Abstract
Introduction
Materials and Methods
Results and Discussion
References
MATERIALS AND METHODS
Top
Abstract
Introduction
Materials and Methods
Results and Discussion
References
1.3, and 1.8°C in 1997, 1998, and 1999, respectively;
the annual levels of precipitation were 1,256, 1,472, and 1,780 mm,
respectively. The annual soil thaw season is very short (between June
or July and September). Summer temperatures vary greatly from near
freezing to more than 20°C at the soil surface (Fig.
1 and Table
1).
View larger version (42K):
[in a new window]
FIG. 1.
Monthly precipitation (bars) and mean air temperatures
(line) at the field study site.
TABLE 1.
Sampling dates and temperatures recorded in air and in
lysimeters during the field study
Soil.
The soil investigated was a mixture of carbonaceous
gravel and sand; the geological underground (C horizon) was central
alpine gneiss. The soil had a pH of 8.0 (measured in 10 mM
CaCl2) and contained 0.014% total N (Kjeldahl), 0.34%
organic C, 2.3% inorganic C, and 19.2% carbonate. The
P2O5 and K2O contents (calcium
lactate extract) and the Fe and Mn contents (EDTA extract) were 20, 430, 580, and 40 µg g of soil1, respectively. The soil
contained 2,612 ± 70 µg of hydrocarbons g (dry weight) of
soil1 at the beginning of the investigation (determined
as described below); the contamination consisted of biodiesel oil.
Field lysimeters (mesocosms).
On 4 July 1997, about 150 kg
of soil was removed from the contaminated zone in the motor pool area
(i.e., in front of the garages and the petrol station for the motor
vehicles used for preparation of ski runs), 50 m north of the
Eisgrat station of the Stubaier Gletscherbahn. The soil was collected
from an approximately 20-m2 area from the surface to a
depth of about 0.1 m. After thorough mixing, about 22 kg of soil
(density, 1.7 g cm3) was placed into each of six
lysimeters. Each lysimeter consisted of two polyethylene pans (length,
0.4 m; width, 0.32 m; fill height, 0.1 m), and one pan
was mounted on top of the other; the bottom of the upper pan, which
contained the soil, had drain holes so that the aqueous soil leachate
could be collected in the lower pan.
1, 41 µg of NO3-N g [dry weight] of soil1, and
41 µg of P g [dry weight] of soil1). One year later,
after the first winter, nutrients were added again at a C/N ratio of
20:1, and the C concentration was related to the measured contamination
(the calculated concentrations were 33 µg of NH3-N g
[dry weight] of soil1, 19 µg of NO3-N g
[dry weight] of soil1, and 19 µg of P g [dry
weight] of soil1).
The lysimeters were transported to an undisturbed area (where visitors
were not allowed) about 50 m south of the soil sampling area and
incubated until 23 September 1999 in the field under natural conditions
(i.e., without any intervention, such as correction of the water
content, etc.). Sampling was done immediately after the lysimeters were
set up and at regular intervals during the summer seasons (Table 1) but
was not possible between October and May, when the soil was frozen. The
soil in the lysimeters was thoroughly mixed, and approximately 700-g
samples, equally distributed, were taken from each pan. The volume of
the soil leachates was recorded; 3 liters of leachate was collected
from each lysimeter. Soil and leachate samples were transported in cooled boxes to the laboratory in order to perform the measurements described below. Each sample from each lysimeter was analyzed in triplicate.
Physical and chemical analyses. Soil dry weight was determined from the weight loss after heat treatment (20 h at 80°C). Soil pH was determined with a glass electrode; 1 part of soil was mixed with 2.5 parts of 10 mM CaCl2. The available inorganic soil nutrient contents were determined as described in detail by Schinner et al. (30). Available nitrogen was extracted from soil by shaking the soil with 2 M KCl for 1 h. The ammonium-N in a filtrate was quantified colorimetrically by measuring the salicylic acid analogue of indophenol blue. The nitrate-N content was determined by measuring UV absorption at 210 nm; a correction for interfering substances was made by reducing nitrate with copper-sheathed, granulated zinc. The available phosphorus was extracted from soil by shaking the soil with 0.4 M LiCl and quantified colorimetrically in the filtrate by the molybdenum blue method.
Total petroleum hydrocarbons (TPH) were measured by the German standard method (10). Ten grams of soil was dehydrated with Na2SO4 and mixed for 30 min with 10 ml of 1,1,2-trichlorotrifluoroethane; the TPH content of the filtrate was quantified by infrared spectroscopy.Soil biological analyses. The following analyses were carried out as described in detail previously (30). Catalase activity was determined by measuring the amount of O2 that evolved from hydrogen peroxide in a phosphate-buffered (pH 6.8) soil suspension. To determine lipase activity, the butyric acid released from tributyrin at 25°C was extracted with ethyl acetate and quantified titrametrically by using 5 mM NaOH. Soil respiration (CO2 evolution) was determined by the Isermeyer technique; the CO2 produced during incubation for 48 h at 10°C was quantified by titration. Soil microbial counts were determined by the plate count method for viable cells (21) on agar plates that contained purified agar and no antifungal agents. R2A agar plates (29) were used to enumerate heterotrophic microorganisms. Hydrocarbon utilizers were quantified on oil agar plates (21) that contained a phosphate-buffered neutral-pH mineral medium and diesel oil as the sole carbon source. CFU were counted after incubation for 14 days at 10°C. No significant growth was observed on control plates without diesel oil.
Soil leachates (hydrocarbon quantification, inhibition of bioluminescence). The TPH contents of soil leachates were determined as described previously (20). Hydrocarbons were extracted from 900 ml of leachate by 10 min of vigorous shaking with 20 ml of 1,1,2-trichlorotrifluoroethane and were quantified by infrared spectroscopy (10).
Inhibition of bioluminescence of Vibrio fischeri (Photobacterium phosphoreum), which is reduced in the presence of toxic compounds (the degree of inhibition depends on the toxicity), was determined by the German standard method (11). The soil leachate was amended with NaCl (2%). Light emission by V. fischeri (BioOrbit, 1243-500 BioTox kit) was measured luminometrically at 15°C immediately and 30 min after various dilutions of the leachates were added. NaCl (2%) was used as the control. Inhibition of bioluminescence was expressed as a GL value, the smallest dilution factor (G) for the test solution (leachate) which resulted in inhibition of light emission of
20%. The test solution
was not toxic at a GL value of 1 or 2, while toxicity was indicated by
GL values greater than 8 (15).
Statistical analyses. Normal distribution of the data was tested by the Kholmogorov-Smirnov test. Whether a treatment had a significant effect on the measured parameters was analyzed by the t test for independent samples (P < 0.05) for data with a normal distribution, while nonparametric data were analyzed by the Mann-Whitney two-sample U test (P < 0.05). Correlations between the measured parameters were analyzed by the Pearson product-moment correlation technique (data with a normal distribution) or the Spearman rank order correlation technique (nonparametric data).
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RESULTS AND DISCUSSION |
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Abstract
Introduction
Materials and Methods
Results and Discussion
References
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Soil hydrocarbon decontamination.
Figure
2A shows
the effect of treatment on the time course of hydrocarbon
disappearance. At the beginning of the study (2,612 ± 70 µg of
hydrocarbons g [dry weight] of soil1), chemical
oxidation processes and metabolic activities, including biodegradation,
were stimulated by manipulation (soil sampling, mixing), and thus the
level of hydrocarbon loss was high with both treatments during the
first 3 weeks. The main difference between treatments occurred during
the next 8 weeks, when biostimulation resulted in significantly
increased hydrocarbon loss. At the end of the first summer season
(after 78 days), a significantly reduced hydrocarbon content,
corresponding to a decontamination level of (42 ± 6)%, was measured
for the fertilized soil (most of the hydrocarbon loss in this soil
occurred during the first summer season), whereas no hydrocarbon loss
was noticed in the unfertilized (i.e., naturally attenuated) soil due
to an apparent significant increase in the hydrocarbon level
accompanied by significant decreases in microbial counts and soil
respiration (Fig. 2C and D) after an initial significant decrease. Such
a release of hydrocarbons has been attributed to enhanced hydrocarbon
mobilization by microbial biosurfactant production (20).
We assume that the mobilized hydrocarbons were biodegraded due to the
favorable nutrient conditions in the fertilized soil but not in the
unfertilized soil.
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1 ([70 ± 2]% reduction), while at the same time
the hydrocarbon concentration was 1,296 ± 110 µg g (dry weight)
of soil1 ([50 ± 4]% reduction) with natural
attenuation. Thus, the level of contamination was still high with both
treatments, but the initial fertilization treatment was an appropriate
treatment in terms of accelerated hydrocarbon loss. Results comparable
to those obtained in our study were obtained in a field study of
bioremediation of aged diesel fuel conducted during two successive
summers in an arctic tundra soil (28).
The most direct way to measure bioremediation efficacy is to monitor
hydrocarbon disappearance rates (4). In our study, the
mean hydrocarbon content in all three summer seasons was significantly lower in the fertilized soil (Table 2). Biostimulation was most effective in the first summer (the rate of hydrocarbon disappearance was 13.9 µg of hydrocarbons g [dry weight] of soil1
day1); however, the positive effect did not last for the
whole study. Despite addition of nutrients after the first winter, the
biostimulation effect decreased. Both in the second summer and in the
third summer natural attenuation resulted in higher rates of
hydrocarbon disappearance (7.0 and 3.8 µg g1
day1, respectively) than of biostimulation (2.7 and 3.3 µg g1 day1, respectively). The natural
attenuation process was slower, but nevertheless it was effective over
a longer period of time.
Several studies have reported favorable effects of fertilization on oil
biodegradation at low temperatures in arctic soils (7, 25, 28,
37), alpine soils (17-19), and antarctic soils (1, 35). An understanding of nutrient effects at a
specific site is essential for successful bioremediation (7, 8,
27). For old contaminations it is not clear that fertilization
has a beneficial effect (13, 20). Hydrocarbon loss is
known to decrease with time. At concentrations below possible threshold concentrations (which may depend on the soil structure and on the
composition of the contaminant), biodegradation rates are low or
negligible (2); this can be attributed to the formation of
persistent polar compounds (24). Residual contamination is obviously greater in old contaminations than in fresh contaminations; in laboratory studies we obtained a level of decontamination of about
90% for experimentally diesel-oil-contaminated soils
(17).
Soil water content and soil pH.
Low moisture content is an
important limiting factor in biodegradation. The soil water content
varied according to the weather and ranged from 1 to 16% at the time
of soil sampling (Table 1). The soil buffering capacity was sufficient
to maintain the soil pH in the neutral range, which is favorable for
biodegradation (4, 27). The initially measured soil pH (pH
8.0 ± 0.06) decreased significantly in the fertilized soil (pH
7.7 ± 0.06) during the first 22 days of the study and then
increased again (pH 7.9 ± 0.1). It was significantly lower (pH
7.5 ± 0.1) after the first winter season, and it decreased
further to pH 7.3 ± 0.06 (unfertilized soil) or pH 7.2 ± 0.1 (fertilized soil) by the end of the second summer season. For both
treatments, a pH of 7.4 ± 0.06 was obtained at the end of the
study. A significantly lower pH in fertilized soil samples than in
corresponding unfertilized soil samples has been described previously
(20). In our study, the effect of biostimulation on soil
pH decreased with time; the pH was significantly decreased by
biostimulation in the summer in 1997 and 1998, while no effect was
observed in the summer in 1999 (Table 2).
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Available nutrients (ammonium-N, nitrate-N, P).
The soil
investigated was nutrient deficient. The extractable available nutrient
contents were around 1 µg g (dry weight) of soil1 and
did not change significantly during the field study (data not shown).
The recommended C/N ratios for soil hydrocarbon bioremediation vary
greatly and range at least from 100:1 to 10:1 (4, 26). It
is important to add both N and P (1, 7, 25), although N
has been shown to be the major limiting nutrient in arctic soils (7). In our study, during the whole study period
fertilization resulted in significantly increased ammonium-N,
nitrate-N, and P contents (Fig. 2B). Nutrient contents decreased
considerably with time. At the end of the first summer period, the
ammonium-N, nitrate-N, and P contents were 23 ± 2, 5 ± 2, and 18 ± 2 µg g (dry weight) of soil1,
respectively. Also, after the second fertilization, which was applied
after the first winter period, the nutrient contents decreased with
time. No significant changes were observed in the third summer. The
disappearance of available nutrients over time can be attributed to
metabolism, immobilization in biomass, immobilization on soil colloids,
and washing out (the latter was confirmed by soil conductivity data
obtained with soil leachates). The nitrate-N content decreased faster
than the ammonium-N content because the level of immobilization of
nitrate-N on soil matrix compounds is low (30). The
decrease in the P content was caused by microbial metabolism and also
by immobilization as apatite (calcium phosphate). The P content
decreased much less than the N content, possibly because of P
remobilization from apatite during microbial metabolism
(14).
Soil leachates.
Aqueous soil leachates correspond to natural
washing out (Table 1). The hydrocarbon content of soil leachates
represents the water-mobilizable components of contamination and is an
important parameter in regulations. During our entire study, the
hydrocarbon content was close to or less than 0.1 mg liter of
leachate1, independent of the treatment. The pH remained
in the neutral range (pH 7.0 to 7.5), and there was no inhibition of
bioluminescence, which is frequently used for toxicity assessment
(5). These results indicate that there was no mobilization
of toxic hydrocarbon fractions.
Soil microbial counts.
At the beginning of our study, we
counted (6.5 ± 0.4) × 107 culturable
heterotrophic microorganisms g (dry weight) of soil1. In
the unfertilized soil, the counts remained almost unchanged in the
summer of 1997 and decreased in the summer of 1998. Biostimulation resulted in a significant increase in the heterotroph counts during the
first summer (the counts increased considerably in July and August and
decreased in September) but did not have a significant effect in the
second and third summer seasons, when the number of heterotrophs
decreased. At the end of the study, the counts were comparable
(1.2 × 107 CFU g [dry weight] of
soil1) for the two treatments (Fig. 2C and Table 2). A
considerable portion of the heterotrophs ([4 ± 0.2] × 106 CFU g [dry weight] of soil1) was able
to utilize diesel oil as a sole carbon source. During the entire study,
the time course for culturable oil degraders was almost comparable to
that for heterotrophs; however, the fluctuations were greater, and the
number of culturable oil degraders was significantly lower than the
number of heterotrophs (Fig. 2C). Biostimulation resulted in
significantly increased counts of oil degraders in the first summer,
while no effect was detected in the second and third summer seasons
(Table 2). At the end of the study, (2.7 ± 1.7) × 106 and (1.5 ± 0.5) × 106 CFU g
(dry weight) of soil1 were present in the unfertilized
and fertilized soil, respectively.
Soil microbial activities. As the use of poisoned controls is precluded in the field, it was of interest to measure the activity of the soil microorganisms in response to treatment. Flat or depressed microbial activity signifies little or no microbial involvement, while strong positive responses indicate that the biodegradative contribution of the indigenous microorganisms is significant (34). Soil respiration and enzyme activities are measures of microbial activity in soil (30) and are indicative of the onset of hydrocarbon biodegradation (21, 34).
The levels of microbial activity in the unfertilized soil fluctuated around the background levels during the whole study. They were at or below the detection limit of the methods used; however, the presence of viable microorganisms was shown by microbial counts. Fertilization resulted immediately in marked but short-lived increases in soil respiration and enzyme activities (Fig. 2D to F). This pattern was also observed in the second and third summer seasons, although the activities were much lower. Soil respiration was significantly enhanced in the first summer season, while catalase and lipase activities were significantly stimulated in all three summer seasons (Table 2). This may have been a side effect of the very low enzyme activities in the unfertilized soil. Over time, there was a tendency for microbial activities to decline to background levels. This decline was associated with the loss of more labile contaminant components due to biodegradation (23). The increases in both microbial counts and activity after the initial fertilization were not repeated to the same extent after refertilization in the second year, despite the fact that our nutrient data (Fig. 2B) do not indicate that conditions were nutrient limiting. Interestingly, we observed that the period when maximum microbial activity occurred in the second summer was not identical to the period when the nutrient level was highest, as was the case in the first year. It also did not correlate with the highest soil surface temperatures (Table 1). The possibility that nitrite toxicity occurred can be excluded since nitrite-N was not detected at any time (data not shown). We can also exclude the possibility that overfertilization occurred, since the nutrient concentrations added were well below concentrations that were found to inhibit microbial activity and hydrocarbon loss in cold soils (25; reference 7 and references therein). It has been suggested that bioremediation is not likely to be effective with extensively degraded oil (8). Aged hydrocarbon residues were not bioavailable to metabolically competent degrading microorganisms (31). We have found that monitoring of soil lipase activity is a valuable indicator of diesel oil biodegradation in freshly contaminated, unfertilized and fertilized soils (22). In artificially contaminated laboratory microcosms, lipase activity remained stable even when the rates of hydrocarbon loss were considerably decreased (21, 22). However, in our field study, lipase activity was negligible in the unfertilized soil and decreased rapidly in the fertilized soil after an initial increase. This pattern was observed in all three summer seasons (Fig. 2F). Interfering factors, such as the composition of the contaminant components and the level of recalcitrant compounds, may influence lipase activity in soils in which the contamination is aged.Correlations.
Correlations between parameters measured in the
field study are presented in Table 3. In
the fertilized soil, all biological parameters (microbial counts, soil
respiration, enzyme activities) correlated significantly positively
with each other, as well as with the residual hydrocarbon
concentration, indicating the importance of biodegradation. The
positive correlations of the available nutrient content with the
hydrocarbon concentration, the microbial counts, and the activities in
the fertilized soil indicate the relevance of nutrients. The low
microbial activities in the unfertilized soil (Fig. 2D to F), as well
as the lack of correlation with the residual hydrocarbon concentration,
led to the conclusion that a considerable part of the hydrocarbon loss
due to natural attenuation probably had to be attributed to abiotic
processes.
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ACKNOWLEDGMENTS |
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This work was supported by the Land Tyrol and by the Austrian Federal Ministries of Science and Environment.
We are especially grateful to H. Klier (Wintersport Tirol AG & Co. Stubaier Bergbahnen KG) for giving us permission to conduct the field experiment. We thank M. Neuner (Amt der Tiroler Landesregierung, Hydrogeographie) for providing the climate data and P. Thurnbichler for technical assistance.
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FOOTNOTES |
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* Corresponding author. Mailing address: Institute of Microbiology, University of Innsbruck, Technikerstrasse 25, A-6020 Innsbruck, Austria. Phone: (43 512) 507-6021. Fax: (43 512) 507-2929. E-mail: rosa.margesin{at}uibk.ac.at.
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REFERENCES |
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Abstract
Introduction
Materials and Methods
Results and Discussion
References
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| 1. | Aislabie, J., M. McLeod, and R. Fraser. 1998. Potential for biodegradation of hydrocarbons in soil from the Ross Dependency, Antarctica. Appl. Microbiol. Biotechnol. 49:210-214. |
| 2. | Allard, A. S., and A. H. Neilson. 1997. Bioremediation of organic waste sites: a critical review of microbiological aspects. Int. Biodeterior. Biodegrad. 39:253-285[CrossRef]. |
| 3. |
Atlas, R. M.
1981.
Microbial degradation of petroleum hydrocarbons: an environmental perspective.
Microbiol. Rev.
45:180-209 |
| 4. | Atlas, R. M., and R. Bartha. 1992. Hydrocarbon biodegradation and oil spill bioremediation. Adv. Microb. Ecol. 12:287-338. |
| 5. | Bitton, G., and B. Koopman. 1992. Bacterial and enzymatic bioassays for toxicity testing in the environment. Rev. Environ. Contam. Toxicol. 125:1-22[Medline]. |
| 6. | Braddock, J. F., J. E. Lindstrom, and E. J. Brown. 1995. Distribution of hydrocarbon-degrading microorganisms in sediments from Prince William Sound, Alaska, following the Exxon-Valdez oil spill. J. Mar. Pollut. Bull. 30:125-132. |
| 7. | Braddock, J. F., M. L. Ruth, J. L. Walworth, and K. A. McCarthy. 1997. Enhancement and inhibition of microbial activity in hydrocarbon-contaminated arctic soils: implications for nutrient-amended bioremediation. Environ. Sci. Technol. 31:2078-2084[CrossRef]. |
| 8. | Bragg, J. R., R. C. Prince, E. J. Harner, and R. M. Atlas. 1994. Effectiveness of bioremediation for the Exxon Valdez oil spill. Nature 368:413-418. |
| 9. | Coulson, S. J., I. D. Hodkinson, A. T. Strathdee, W. Block, N. R. Webb, J. S. Bale, and M. R. Worland. 1995. Thermal environments of Arctic soil organisms during winter. Arct. Alp. Res. 27:364-370[CrossRef]. |
| 10. | Deutsches Institut für Normung e.V.. 1981. DIN 38 409-H18. Deutsche Einheitsverfahren zur Wasser-, Abwasser- and Schlammuntersuchung. Bestimmung von Kohlenwasserstoffen (H18). Deutsches Institut für Normung e.V., Berlin, Germany. |
| 11. | Deutsches Institut für Normung e.V.. 1991. DIN 38412-L34. Testverfahren mit Wasserorganismen (Gruppe L); Bestimmung der Hemmwirkung von Abwasser auf die Lichtemission von Photobacterium phosphoreum-Leuchtbakterien-Abwassertest mit konservierten Bakterien. Deutsches Institut für Normung e.V., Berlin, Germany. |
| 12. | Gounot, A. M. 1999. Microbial life in permanently cold soils, p. 3-15. In R. Margesin, and F. Schinner (ed.), Cold-adapted organisms. Springer, Berlin, Germany. |
| 13. | Hinchee, R. E. 1998. In situ bioremediation: practices and challenges, p. 17-20. In R. Serra (ed.), Biotechnology for soil remediation. CIPA, Milan, Italy. |
| 14. |
Illmer, P., and F. Schinner.
1995.
Solubilization of inorganic calcium phosphates solubilization mechanisms.
Soil Biol. Biochem.
27:257-263[CrossRef].
|
| 15. | Kreysa, G., and J. Wiesner (ed.). 1995. Bioassays for soils. Dechema e.V., Frankfurt, Germany. |
| 16. |
Lindstrom, J. E.,
R. C. Prince,
J. C. Clark,
M. J. Grossmann,
T. R. Yeager,
J. F. Braddock, and E. J. Brown.
1991.
Microbial populations and hydrocarbon biodegradation potentials in fertilized shoreline sediments affected by the T/V Exxon Valdez oil spill.
Appl. Environ. Microbiol.
57:2514-2522 |
| 17. | Margesin, R., and F. Schinner. 1997. Bioremediation of diesel-oil contaminated alpine soils at low temperatures. Appl. Microbiol. Biotechnol. 47:462-468[CrossRef]. |
| 18. | Margesin, R., and F. Schinner. 1997. Efficiency of indigenous and inoculated cold-adapted soil microorganisms for biodegradation of diesel oil in alpine soils. Appl. Environ. Microbiol. 63:2660-2664[Abstract]. |
| 19. | Margesin, R., and F. Schinner. 1999. Biological decontamination of oil spills in cold environments. J. Chem. Technol. Biotechnol. 74:1-9. |
| 20. | Margesin, R., and F. Schinner. 1999. A feasibility study for the in situ remediation of a former tank farm. World J. Microbiol. Biotechnol. 15:615-622[CrossRef]. |
| 21. | Margesin, R., G. Walder, and F. Schinner. 2000. The impact of hydrocarbon remediation (diesel oil and polycyclic aromatic hydrocarbons) on enzyme activities and microbial properties of soil. Acta Biotechnol. 20:313-333[CrossRef]. |
| 22. |
Margesin, R.,
A. Zimmerbauer, and F. Schinner.
1999.
Soil lipase activitya useful indicator of oil biodegradation.
Biotechnol. Tech.
13:859-863[CrossRef].
|
| 23. | McGill, W. B., M. J. Rowell, and D. W. S. Westlake. 1981. Biochemistry, ecology, and microbiology of petroleum components in soil, p. 229-296. In E. A. Paul, and J. N. Ladd (ed.), Soil biochemistry, vol. 5. Marcel Dekker, Inc, New York, N.Y. |
| 24. | Miethe, D., V. Riis, and W. Babel. 1996. Zum Problem der Restkonzentration beim mikrobiellen Abbau von Mineralölen. Hamburger Ber. 10:289-302. |
| 25. | Mohn, W. W., and G. R. Stewart. 2000. Limiting factors for hydrocarbon biodegradation at low temperature in Arctic soils. Soil Biol. Biochem. 32:1161-1172[CrossRef]. |
| 26. | Morgan, P., and R. J. Watkinson. 1989. Hydrocarbon degradation in soils and methods for soil biotreatment. Crit. Rev. Biotechnol. 8:305-333[Medline]. |
| 27. | Norris, R. D. 1994. Handbook of bioremediation. CRC Press, Boca Raton, Fla. |
| 28. |
Piotrowski, M. R., and R. G. Aaserude.
1992.
Bioremediation of diesel contaminated soil and tundra in an Arctic environment, p. 115-141.
In
P. T. Kostecki, and E. J. Calabrese (ed.), Contaminated soilsdiesel fuel contamination. Lewis Publishers Inc., Chelsea, Mich.
|
| 29. |
Reasoner, D. J., and E. E. Geldreich.
1985.
A new medium for the enumeration and subculture of bacteria from potable water.
Appl. Environ. Microbiol.
49:1-7 |
| 30. | Schinner, F., R. Öhlinger, E. Kandeler, and R. Margesin (ed.). 1996. Methods in soil biology. Springer, Berlin, Germany. |
| 31. | Smith, M. J., G. Lethbridge, and R. G. Burns. 1999. Fate of phenanthrene, pyrene and benzo(a)pyrene during biodegradation of crude oil added to two soils. FEMS Microbiol. Lett. 173:445-452[CrossRef][Medline]. |
| 32. |
Song, H. G., and R. Bartha.
1990.
Effects of jet fuel on the microbial community of soil.
Appl. Environ. Microbiol.
56:646-651 |
| 33. | U.S. Environmental Protection Agency. 1999. Use of monitored natural attenuation at Superfund RCRA corrective action, and underground storage tank sites. OSWER Directive 9200.4-17. U.S. Environmental Protection Agency, Washington, D.C. |
| 34. | Wang, X., and R. Bartha. 1990. Effects of bioremediation on residues, activity and toxicity in soil contaminated by fuel spills. Soil Biol. Biochem. 22:501-505[CrossRef]. |
| 35. | Wardell, L. J. 1995. Potential for bioremediation of fuel-contaminated soil in Antarctica. J. Soil Contam. 4:111-121. |
| 36. |
Whyte, L. G.,
J. Hawari,
E. Zhou,
L. Bourbonnière,
W. E. Inniss, and C. W. Greer.
1998.
Biodegradation of variable-chain-length alkanes at low temperatures by a psychrotrophic Rhodococcus sp.
Appl. Environ. Microbiol.
64:2578-2584 |
| 37. | Whyte, L. G., L. Bourbonnière, C. Bellerose, and C. W. Greer. 1999. Bioremediation assessment of hydrocarbon-contaminated soils from the High Arctic. Bioremed. J. 3:69-79. |
Applied and Environmental Microbiology, July 2001, p. 3127-3133, Vol. 67, No. 7
0099-2240/01/$04.00+0 DOI: 10.1128/AEM.67.7.3127-3133.2001
Copyright © 2001, American Society for Microbiology. All rights reserved.
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