Biodegradation of aged residues of atrazine and alachlor in a mix-load site soil

Soil Biology & Biochemistry 41 (2009) 2484–2492

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Biodegradation of aged residues of atrazine and alachlor in a mix-load site soil Anastasia E.M. Chirnside*, William F. Ritter, Mark Radosevich University of Delaware, Bioresources Engineering, 531 South College Ave, Newark, DE 19716-2140, United States

a r t i c l e i n f o

a b s t r a c t

Article history: Received 26 March 2009 Received in revised form 28 August 2009 Accepted 3 September 2009 Available online 23 September 2009

A selected microbial consortium (SMC) capable of degrading two specific herbicides, alachlor (2-chloro20 ,60 -diethyl-N-[methoxymethyl]-acetanilide; AL) and atrazine (2-chloro-4-ethylamino-6-isopropylaminoS-triazine; AT) was isolated from a pesticide-contaminated mix-load site soil. Evaluation of bioaugmentation as a feasible bioremediation strategy for this mix-load site soil (Site 5A) was initiated in standard laboratory biometer flasks utilizing the isolated SMC. The biometer flasks were monitored for CO2 evolution and pesticide degradation. The total amount of CO2 evolved from the treated biometer flasks was significantly different from the control flasks. The rate of CO2 evolution was 2.6 times faster in the treated soil (0.0123 mM CO2 d1 vs. 0.0048 mM CO2 d1). The total net CO2 produced in the treated biometer flasks was 0.9481 mM, representing mineralization of approximately 10% of the AT and AL initially present. Forty-eight percent of AT and 70% of AL was degraded in the inoculated biometer flasks. The first-order rate constants were 0.0064 d1 and 0.1331 d1 for AT and AL, respectively. The calculated half-life of AT was 108 d while a 50% decrease in AL occurred by Day 5. In just 2 d, 20% of the AT was degraded while only 10% of the AL disappeared. The initial fast degradation rate of AT was followed by a much slower, more gradual degradation rate period that lasted about 35 d. Alternatively, the rate of AL degradation increased after the second day resulting in 60% of the AL being transformed by the end of the first week. Alachlor degradation appeared to be dependent upon AT degradation especially during the first several days of the incubation period. Complete disappearance of the herbicides over the study time was not achieved. Ó 2009 Elsevier Ltd. All rights reserved.

Keywords: Atrazine Alachlor Selected microbial consortium Bioaugmentation Degradation rate Half-life Pesticide mix-load site

1. Introduction Many agricultural dealerships’ and pesticide mix-load sites’ soils and associated groundwater have contained such extremely high concentrations of pesticides, fertilizers and other organic compounds used in pesticide formulations that they have been designated as point sources of contamination (Ames and Hoyle, 1999; and Habecker, 1989). Remediation of these sites has been problematic due to the complex mixture of contaminates present, and due to the natural heterogeneity of the soil and water environment (Ames and Hoyle, 1999). Bioremediation has been examined as a cleanup technology because it is thought to be a low cost alternative to chemical treatment procedures. However, bioremediation treatment strategies must still evaluate and overcome the difficulties associated with mix-load sites. The utilization of bioaugmentation, the addition of specialized microbes with enhanced capabilities of degrading the compound(s) of interest, has proven to be ineffective in many cases because evaluation of the actual field

* Corresponding author. Tel.: þ1 302 831 8871; fax: þ1 302 831 2469. E-mail address: [email protected] (A.E.M. Chirnside). 0038-0717/$ – see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.soilbio.2009.09.005

conditions was not complete (Schmidt and Scow, 1996). Important information to consider in site characterization is the concentration of the compounds present; the presence of mixtures of compound types; the presence of any non-aqueous phase liquids; the physical quality of the solid, liquid and gas phases; the biodiversity of the site and contaminant transport and fate. Development of the selected microbial consortium (SMC) has usually concentrated only on one or two of the contaminants actually present at a mix-load site. Therefore, inoculum successful in laboratory studies may fail in the field because of sensitivity of the specialized microbes to high concentrations of non-target compounds (Edgehill, 1999; and Scalzi et al., 2001). An SMC capable of degrading the herbicides AT and AL was isolated from a contaminated soil from a 100-year-old mix-load site located in Reading, PA. The SMC was selected on its ability to degrade AT and AL, both present in the mix-load site soil (Chirnside et al., 2007). However, the mix-load site soil also contained cyanazine {2-[(4-chloro-6-ethylamino-s-triazine-2-yl)amino]-2methylpropionitrile; CY} and metolachlor (2-chloro-N-[2-ethyl-6methyl-phenyl]-N-[2-methoxy-1-methyl-ethyl]acetamide; ME) at concentrations higher than the target compounds, AT and AL. The presence of additional substrates may possibly inhibit degradation

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of the AT and AL through catabolic repression. Structurally related compounds may share comparable degradation pathways, which may result in the buildup of the same metabolites. For example, deisopropylatrazine (CEAT) is a metabolite of AT and CY degradation. An elevated concentration of CEAT may or may not be detrimental to AT degradation. Most of the studies on AT mineralization have shown that CEAT is quickly degraded and not considered the ratelimiting step of the degradative pathway (Radosevich et al., 1995). Ostrofsky et al. (2001) found that adding cyanuric acid, the central metabolite in s-triazine degradation, had no effect on the mineralization rate. Because dealkylation of s-triazines can be coupled with microbial C and energy metabolism (Crawford et al., 2000), the dealkylation of the additional substrates could increase the number of AT degrading microorganisms resulting in greater degradation. Alternatively, these other substrates may decrease the degradation of AT and/or AL through competitive inhibition where the microbes selectively degrade the additional substrates before the target compounds. Presence of CY may reduce the effectiveness of the SMC in field trials to degrade AT. Cyanazine has been shown to be more easily degraded than AT (Mandelbaum et al., 1995; Meyer et al., 2001). The presence of ME, an acetanilide herbicide similar to AL, could also have a deleterious effect on the SMC’s ability to degrade AL. However, ME is reported to be more persistent within the soil than AL (Cookson, 1995). The mixture of contaminates in the mix-load site soil may also have some positive influences on the degradative ability of the isolated SMC. For instance, degradation of the more readily transformed CY could provide C and energy for growth resulting in a larger population of triazine degraders. This would ultimately result in greater degradation of AT. Many degradation schemes are possible especially when a consortium of microorganisms is utilized for degradation. Presence of the additional substrates could initiate cometabolism of the desired compounds. Both Pseudomonas spp. and Alcaligenes spp. have been shown to cometabolize many substrates (Alexander, 1994 and Cookson, 1995). The SMC utilized in this study contains both of these genera. Mixed microbial populations have a greater possibility of degrading recalcitrant compounds because of their larger catabolic gene pool (Fourner et al., 1997). Therefore, in order to understand microbial interactions and to evaluate the field success of the SMC, laboratory degradation experiments were performed with the actual mix-load site contaminated soil. The specific objective was to evaluate the ability of the isolated SMC to degrade AT and AL within the mix-load site soil and to determine the reaction pathways and degradation rates involved. 2. Materials and methods

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a side-arm trap. Fifty grams of air-dried, ground soil from Site 5A (Table 1) were placed into the main body of the biometer flasks. Site 5A soil was chosen for this study because of the lack of AT and AL degraders (Chirnside et al., 2007) and the presence of high concentrations of the herbicides, AT, AL, ME and CY (Table 1). The flasks were autoclaved at 120  C (20 psi) on three consecutive days; 60 min the first day and then 30 min for the next 2 d. The side-arm trap was filled with 10 mL of 0.2 N NaOH (reagent grade, VWRSP) prepared with CO2-free deionized distilled water (DDW). The main body of the biometer was capped with a rubber stopper containing a covered stopcock filter reservoir filled with Drierite (W.A. Hammond). The trap was capped with a rubber stopper containing a Luer needle tip that was also closed with a stopper. Six biometer flasks were utilized for the experiments. The 3 control flasks received 5 mL of sterile DDW and the 3 treatment flasks received 5 mL of the SMC inoculum resulting in a soil cell density of 1.10  0.55  107 cells g1 soil. After addition of the solutions, the flasks were mixed well with sterile spoons to distribute the inoculum and rewet the soil to approximately 100% of field water holding capacity. Biometer flasks were incubated at 23  C for the length of the experiment. The NaOH in the side traps was collected and analyzed for CO2 concentration while the flasks were aerated through the Drierite filter every 3–4 d. The extracted NaOH was analyzed for CO2 concentration following the traditional acid titration method outlined by Zibilske (1994). The side-arm was carefully recharged with fresh NaOH after each of the sampling events. The biometers were monitored for AT, AL, CY, ME, and another s-triazine, simazine [2-chloro-4,6-bis(ethlyamino) s-triazine,{SI}] periodically. After termination of the incubation period, culturable AT and AL degrading bacteria were enumerated by dilution plate count technique. The entire experiment was done in triplicate. 2.3. Pesticide analysis During the incubation study, the biometer flasks were sampled periodically for herbicide concentration. Sampling was carried out at the same time that the flasks were sampled for CO2. The soils were analyzed for AT and AL using a solid phase extraction method (Chirnside and Ritter, 1992) followed by analysis by high performance liquid chromatography (HPLC). This method was able to detect the herbicides CY, ME and SI as well as AT and AL. Because of the high concentrations of these other herbicides in the biometer soil, it was necessary to monitor their disappearance from the biometer flasks too. The solid phase extraction method was modified in order to extract possible polar metabolites. Aliquots of the biometer flask soil were taken using a sterile spoon on the

2.1. Inoculum preparation The inoculum was prepared from the SMC that was grown up in a buffered, saturated solution of AT and AL and stored at 4  C (Chirnside et al., 2007). A 100 mL aliquot of the cell solution was taken from the SMC growth culture and centrifuged for 15 min at 12,000  g rpm. The supernatant was discarded and the pellets washed with sterile phosphate buffered saline (PBS) solution. The resulting cell solution was centrifuged and re-suspended in sterile PBS two more times. The microbial density of the resulting cell suspension was estimated by both plate count and by dry cell gravimetric methods. 2.2. Biometer flask experimental design Aerobic microcosms were prepared in 250 mL biometer flasks (Bellco, Glass Inc., Vineland, NJ) consisting of a main body with

Table 1 Physical and chemical properties of Site 5A soil (dry wt. basis). Sample SITE S-5A Location Textural class SOM % pH NH4–N (mg kg1) NO3–N (mg kg1) Sol salts (mmho cm1) Acidity (meq 100 g1) Moisture content, field capacity (%) Atrazine (mg kg1) Alachlor (mg kg1) Cyanazine (mg kg1) Simazine (mg kg1) Metolachlor (mg kg1) a

East side of loading dock Loamy sand 1.1 (0)a 7.33 10.49 (0.31) 50.24 (3.12) 1.34 (0.11) 0.033 (0.029) 8.61 205.1 (10.2) 108.5 (8.3) 2272.3 (145.5) 13.15 (1.94) 1829.4 (106.4)

Number in parentheses represents standard deviation of 3 values.

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following days of the incubation period: 0, 3, 5,7,14,21,28,37,118, and 133. During sampling the soil in the flask was vigorously stirred before removing the aliquot. The aliquot of soil was transferred to a tarred Erlenmeyer and weighed. Approximately 1-g samples were removed from the flasks at each sampling date. The soils were extracted by shaking with 50 mL 90% methanol (ultra-pure, HPLC grade; VWRSP) on a mechanical shaker for 3 h and then the extract was then centrifuged at 2400  g rpm for 20 min. A 5 mL aliquot of the supernatant was transferred to a graduated test tube and diluted with 55 mL DDW so that the percentage of methanol in solution was 7.5%, ensuring adequate loading of the compounds of interest. Solid-phase extraction of the diluted aliquots was performed on a semiautomatic vacuum manifold capable of 24 simultaneous extractions. After evaluation of a variety of solid phases, BAKERBOND PolarPlus C18 (octadecyl) solid phase columns (Mallinckrodt Baker Inc., Phillipsburg, NJ) were used for the extractions. The 6 mL disposable columns were packed with 500 mg reversed phase octadecylsilane bonded to silica gel. Preconditioning and activation were achieved by passing 5 mL of the following solvents through the column in succession without letting the column dry out between additions: methanol, methylene chloride, ethyl acetate, methanol, 7.5% methanol and finally DDW. The diluted sample was loaded onto the conditioned column at a rate of 3.0–4.0 mL min1. When sample loading was completed, the cartridges were airaspirated for 15 min to allow the sorbent to dry. All solvents were ultra-pure, HPLC grade (VWRSP). The loaded dry columns were eluted with 5 mL of methanol, which were drawn through to the column until saturation, allowed to sit for 5 min and then eluted at a rate of 1–2 mL min1. This step was repeated with 5 mL of ethyl acetate and then 5 mL of methylene chloride. The 15 mL of solvent was eluted directly into glass test tubes, and then evaporated to dryness in a warm water bath (40  C) with a gentle flow of ultra high-purity nitrogen gas (Keen Gas, Newark, DE). The dried elutant was brought up to 5 mL with acetonitrile, mixed gently and transferred to HPLC vials for analysis. Table 2 lists the herbicides and the metabolites used in this research. The compounds were purchased from Chem Service, Inc.; West Chester, PA and used as received. Also listed in Table 2 is the initial stock concentration of each compound and the solvent utilized for dissolution. All compounds were analyzed individually for method development and then combined into three separate stock solutions for ease of analysis on the HPLC. Percent recoveries were determined by replicate analysis of reference soil spikes. The HPLC system was a Thermo Separations Product-AS3000 Series, West Palm Beach, FL equipped with a UV photodiode array detector

set at 220 nm. The compounds were separated by reversed phase HPLC using a PhenomenexÒ Synergi 4m Polar-RP column, 150  4.6 mm, dp ¼ 4 mm (Phenomenex; Torrance, CA) at ambient temperature and a constant flow rate of 1.4 mL min1. A sample volume of 20 mL was injected via an autosampler. A gradient mobile phase consisting of A) acetonitrile and B) 0.003 M potassium dihydrogen phosphate, pH ¼ 3.00 was used. 3. Results 3.1. Inoculum characterization The inoculum cell densities enumerated by the two methods were within the same order of magnitude. The population density determined by plate counts was 1.10  0.25  108 CFU mL1. Gravimetric determination resulted in 4.9  5  108 cells mL1. 3.2. CO2 evolution rates Fig. 1 shows the average CO2 evolution from both the treatment and the control biometers during the 133-day incubation. A gradual increase in CO2 evolution was seen for both treatments. However, the CO2 evolution in the treated biometers increased at a rate 2.6 times greater than the background CO2 of the untreated biometers (0.0123 mM CO2 d1 vs. 0.0048 mM d1). The total amount of CO2 evolved from the treatment biometers was significantly different from the treatment biometers according to the Kruskal–Wallis oneway analysis of variance on ranks (P < 0.05). The total amount of CO2 evolved from the autoclaved control biometers was only 6% of the theoretical CO2 concentration. Thus, the net CO2 produced in the treatment biometers can be attributed to biological activity (Sharabi and Bartha, 1993). There was a burst in CO2 evolution towards the end of the incubation period at Day 108. At this time, production again decreased approximately 50% to a steady rate to the end of the incubation period. 3.3. Herbicide degradation During the incubation period, atrazine concentration decreased from an initial concentration of 0.722  0.047 mM to a final concentration of 0.410  0.032 mM. A 20% decrease in AT concentration was seen by the second day of the incubation period. As seen in Fig. 2A, AT degradation continued until about Day 14, at which point there was a very slow, steady rate of AT disappearance. The rate of AT degradation increased again after Day 35 and remained constant until the end of the incubation period. No

Table 2 A list of the compounds monitored during experiments. The concentration (mg L1) of the initial stock solution, the solvent used to make the standard, and % recovery at the 95% level of confidence. Compound

Label

Purity

Stock I mg L1

Solvent

% Recovery

Atrazine Triazine deethyl Atrazine deethyl-2-hydroxy Atrazine-2-hydroxy Atrazine deethyl deisopropyl-2-hydroxy Atrazine deethyl deisopropyl Atrazine deisopropyl Alachlor 2-Chloro-20 ,60 -diethylacetamilide 2,6-Diethylanaline Cyanazine Metolachlor Simazine Simazine hydroxy Aniline

CIET CIAT OIAT OIET OAAT CAAT CEAT AL DMA DIE CY ME SI SIOH ANI

99% 99% 98% 98% 99% 96% 97% 99% 98.7% 98.2% 99% 96.5% 99% 99% 99%

100 100 25 25 10 10 10 100 100 100 100 100 100 10 100

Acetonitrile Methanol Methanol Methanol Methanol Acetonitrile Acetonitrile Acetonitrile Acetonitrile Acetonitrile Acetonitrile Acetonitrile Acetonitrile Methanol Acetonitrile

100.5 38.2 8.04 8.00 23.8 24.8 16.3 97.1 100.6 50.1 87.6 84.6 86.4

            

5.8 3.2 2.4 1.7 7.9 6.4 1.8 8.7 9.2 7.4 5.4 6.5 7.6

44.02  11

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CO2 RECOVERED mM

CUMULATIVE CO2 EVOLUTION 2.00 1.75 1.50 1.25 1.00 0.75 0.50 0.25 0.00

y = 0.0123x + 0.1659 R² = 0.9778

Inoculated

Control y = 0.0048x + 0.173 R² = 0.9513

0

10

20

30

40

50 60

70

80 90 100 110 120 130 140 150 160 DAY

Fig. 1. Cumulative CO2 evolution recovered from both control and treatment biometers of the SMC degradation study. Data are means  standard deviation for triplicate experiments. Differences between means with different subscripts are statistically significant according to the Kruskal–Wallis one-way analysis of variance based on ranks (P < 0.05).

degradation was seen in the control biometers. Assuming a pseudofirst order reaction for the disappearance of AT, the plot of the natural logarithm of initialized AT concentration (C/C0) versus time yielded the first-order rate constant of 0.0024 d1 (Fig. 2B). The half-life (t1/2) for AT calculated from the plot was 289 d. During the incubation period, alachlor concentration decreased from an initial concentration of 0.755  0.029 mM to a final concentration of 0.254  0.021 mM. Compared to AT, the initial disappearance of AL proceeded at a slower rate. Only a 10% decrease in AL concentration was seen by the second day of the incubation period. However, Fig. 3A showed that after one week over 60% of the AL had degraded. AL degradation continued until about Day 21, at which point, no further AL degradation was seen. No significant degradation was seen in the control biometers. Again, assuming a pseudo-first order reaction for the disappearance of AL, a plot of the natural logarithm of initialized AL concentration (C/C0) versus time resulted in a first-order rate constant equal to 0.0814 d1 (Fig. 3B). The t1/2 for AL calculated from the plot was 8.5 d.

1.2

a

1.2

1

1

0.9

0.8

0.8 0.7 INOCULATED

0.5 0

3

5

CONTROL

7

14

21

28

37

118

0.2

133

INOCULATED

DAY

0.1

DAY

b LN (C/Co)

-0.1 -0.2 -0.3 -0.4 -0.5 -0.6 0

CONTROL

0

0

LN (C/Co)

0.6 0.4

0.6

0.2

ALACHLOR SMC DEGRADATION

a

1.4

ATRAZINE SMC DEGRADATION

C /Co

C/Co

1.1

Despite the poor extraction recoveries for the AT metabolite compounds monitored during incubation, all but atrazine deethyl (CIAT) was seen in the treated biometer flasks (Fig. 4a, Table 3). During the first few days of the incubation period there was an increase in atrazine deethyl-2-hydroxy (OIAT) that continued until Day 7. By Day 5, a slight increase in atrazine deethyl deisopropyl (CAAT), atrazine-2-hydroxy (OIET) and in atrazine deethyl desisopropyl-2-hydroxy (OAAT) was seen. Except for OIET and OIAT, the metabolites exhibited a similar pattern of production and degradation. After the initial increase, these metabolites gradually decreased from Day 7 until Day 28 at which time a gradual increase was seen. Small amounts of these metabolites were present through the entire incubation period. Conversely, after the initial increase in OIET, it disappears permanently from the inoculated cultures by Day 21. Atrazine-2-hydroxy increased between Day 14 through Day 21 and then exhibited another large increase towards the

3

5

7

14

21

28

37

118

133

b

0.2 0 -0.2 -0.4 -0.6 -0.8 -1 -1.2 0

1

2

3

y = -0.0444x - 0.0651 R2 = 0.8183

k= -0.0444 d-1

4

5

6

7

8

DAY

DAY t1/2= 10.65

Fig. 2. Disappearance of AT over time from biometers of the SMC degradation study. Error bar represents  1 standard deviation of triplicate experiments. A) AT concentration expressed as concentration/initial concentration (C/C0). B) Semi-logarithmic plot of initialized AT concentration.

y = -0.1458x

R2 = 0.8583

k= 0.1458 d -1

t1/2=4.75 d

Fig. 3. Disappearance of AL over time from biometers of the SMC degradation study. Error bar represents  1 standard deviation of triplicate experiments. A) AL concentration expressed as concentration/initial concentration (C/C0). B) Semi-logarithmic plot of initialized AL concentration.

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55% of the SI was transformed (Fig. 5a). Simazine degradation resulted in the loss of 0.0256 mM of SI. The initial concentration of ME was quite high at 1606 mg kg1. Twenty percent of the ME present in the contaminated soil was degraded, but at a much slower rate than the AL (Fig. 5b). 4. Discussion 4.1. Disappearance of atrazine and alachlor

Fig. 4. a) Production of AT metabolites over time in the SMC degradation study. OIET ¼ atrazine-2-hydroxy, CAAT ¼ atrazine deethyl deisopropyl, OAAT ¼ atrazine deethyl deisopropyl-2-hydroxy, CEAT ¼ atrazine deisopropyl, OIAT ¼ atrazine deethyl2-hydroxy. b) Production of AL metabolites over time in the SMC degradation study. DMA ¼ 2-chloro-20 ,60 -diethylacetamilide, DIE ¼ 2,6-diethylanaline.

end of the incubation period coinciding with the flush of net CO2 evolution. Alachlor degradation resulted in the formation of 2-chloro-20 ,60 diethylacetanilide (DMA) and 2,6-diethyl aniline (DIE) (Fig. 4b). However, no aniline was detected throughout the incubation period. Initially, there was an increase in DMA followed by an increase in DIE. However, there was a greater increase in DIE, which continued throughout the incubation period. As seen with atrazine2-hydroxy, DIE exhibited an increase towards the end of the incubation period coinciding with the flush of net CO2 evolution. During the incubation period, 16% of the CY present was degraded (Fig. 5a). The initial CY concentration was very high at 1794 mg kg1 (7.45 mM), thus 1.192 mM of CY was removed. By Day 14, 32% of the SI was transformed and at the end of the incubation Table 3 Overview of AT and AL transformation and of the production and further transformation of metabolites within the SMC treated soil. % AT removed 48.1 Metabolite CEAT CAAT OIET OIAT OAAT Total

mM AT removed 0.3806 Formed (mM) 0.070 0.064 0.485 0.106 0.078 0.803

Degraded (mM) 0.055 0.035 0.408 0.136 0.096 0.730

% AL removed 69.5 Metabolite DMA DIE Total

mM AL removed 0.5269 Formed (mM) 0.205 0.250 0.455

Degraded (mM) 0.142 0.158 0.300

NET CO2 evolution (mM) 0.948

Inoculation of the pesticide-contaminated mix-load site soil with the isolated indigenous SMC resulted in the degradation of both AT and AL. However, complete removal of these herbicides from the soil was not accomplished (Table 3). Initial concentrations of atrazine and alachlor in the contaminated soil were 160 mg kg1 (0.791 mM) and 200 mg kg1 (0.749 mM), respectively. Approximately 48% of the AT (0.381 mM) and 70% of the AL (0.527 mM) were transformed within the biometer flasks during the 133 d incubation period. Based on the net CO2 evolved from the inoculated biometer flasks (0.948 mM CO2), biological activity occurred steadily throughout the incubation period. The significant difference between the inoculated and the autoclaved biometers suggested that minimal mineralization might have occurred. The net CO2 evolved (0.9481 mM) represents 31.1% of the total CO2 evolved if AT was completely mineralized and only 9.1% of the total CO2 evolved if the AL was completely mineralized as well. Complete mineralization of AL has never been reported in the literature (Mangiapan et al., 1997). However, some researchers have measured mineralization to a limited degree. About 1.8% of AL was mineralized in a sandy loam soil (Novick et al., 1986). An isolated a microbial consortium was able to mineralize 12% of the AL present as the sole C and energy (Sun et al., 1990). Therefore, assuming 5% AL mineralization, approximately 10% of the AT and AL initially present was mineralized. This is consistent with the literature. A microbial consortium containing a Pseudomonas sp. was able to mineralize 10% of the AT present in the soil (0.464 mM) (Mandelbaum et al., 1993). Several Pseudomonas spp. have been isolated that are able to mineralize AT (deSouza et al., 1998; Mandelbaum et al., 1995; Yanze-Kontchou and Gschwind, 1994). Analysis of the isolated SMC utilizing the FAME technique identified Pseudomonas sp. (similarity index of 0.839) as one of the single isolates within the SMC (Chirnside et al., 2007). The concentration of SMC organisms remained steady with no significant change in population during incubation indicative of cometabolic biodegradation. Both Pseudomonas spp. and Alcaligenes spp. are known to cometabolize recalcitrant compounds. Genera of these microorganisms were also identified within the SMC. Cometabolism usually exhibited slower degradation rates. The rate constant for AT and AL biodegradation was determined to be 0.0064 d1 and 0.0814 d1, respectively. This degradation rate was lower than others reported in the literature for atrazine degradation in soil, e.g. k ¼ 0.0540 to 0.063, 0.026 d1 (Stolpe and Shea, 1995). Mineralization of atrazine in soil collected from a continuous-corn crop field that was amended with atrazine resulted in first-order rate constants of 0.04–0.14 d1 and half-lives of 5–19 d (Ostrofsky et al., 2001). The degradation rate for AL (0.081 d1) was similar to other values reported in the literature, e.g. k ¼ 0.0798–0.0806 d1 (Stolpe and Shea, 1995). The slower degradation rate constant for AT found in this work could also have been influenced by the nature of the mix-load site soil, which contained aged residues of AT, AL, other herbicides, and inorganic nitrogen at high concentrations. Adding inorganic N suppressed the biodegradation and/or the mineralization of AT in surface soils. Feakin et al. (1994) isolated bacterial strains from soil and

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ALACHLOR & ATRAZINE DEGRADATION

CONCENTRATION mM

1

AL

AT

0.75 0.5 0.25 0

0

3

5

7

14

21

28

37

118

133

DAY Fig. 6. Disappearance of AT and AL over time from biometers of the SMC degradation study. Error bar represents  1 standard deviation of triplicate experiments.

Fig. 5. a) Disappearance of CY and SI over time from biometers of the SMC degradation study. CY-INC ¼ cyanazine concentration in the inoculated treatment, CYCRL ¼ cyanazine concentration in the control (not inoculated), SI-INC ¼ simazine concentration in the inoculated treatment, and SI-CRL ¼ simazine concentration in the control. b) Disappearance of ME over time from biometers of the SMC degradation study. Error bar represents  1 standard deviation of triplicate experiments. Error bar represents  1 standard deviation of triplicate experiments.

wastewater that did not degrade AT in the presence of added N (35 mg L1). Ames and Hoyle (1999) found that added N decreased AT degradation in static cultures with the degradation rate decreasing from 0.0142 d1 to 0.0068 d1, similar to the AT degradation rate measured in this study. An SMC isolated from a mix-load site in Minnesota was able to degrade AT as sole N source (Mandelbaum et al., 1993). When AT was removed from the growth media, no increase in population occurred. Adding NH4NO3 increased growth of the microbes but suppressed AT degradation. The bioavailability of the aged residues may also have been one of the rate-limiting factors for transformation of AT and AL. Bioavailability of a compound to the microorganism is often the rate-limiting step in biodegradation. Old polluted sites, like the mix-load facility, often contain contaminants that appear to be inaccessible for biodegradation (Bosma et al., 1997). When compounds are unavailable to the microbes, they survive with minimum growth even at high contaminant concentrations. Thus, the rate and extent of biodegradation of these contaminants are reduced due to the adsorption of these aged residues. Rates of substrate degradation and subsequent microbial growth are strictly dependent on soil solution concentrations (Shelton and Doherty, 1997). Alachlor concentration decreased at a slower rate than AT during the first few days of the incubation and it disappearance appeared to be dependent upon AT degradation (Fig. 6). By Day 3, 22.1% of AT was degraded while only 10.8% of AL, which gives a ratio of approximately 2 mM of AT to 1 mM AL. But by Day 5, the rate of AL degradation increased while AT degradation slowed down

considerably (25.1% AT vs. 41.7% AL). The ratio of the amount of AT degradation to the amount of AL dropped to 0.602:1. One of the first steps in AT degradation has been considered to be dealkylation and deamination of the side chains of the s-triazine ring (Yassir et al., 1999). The removal of these side chains from the AT molecule have been coupled to microbial C and energy metabolism and NH4–N assimilation (Crawford et al., 2000; Strong et al., 2002). It is possible that the energy and N derived from the side-chain removal enabled the consortium to further degrade the C rich, N poor AL moiety. With the additional carbon of AL, the consortium was able to further degrade the triazine ring to access the N’s. Application of C after formation of the dealkylated metabolites increased AT mineralization (Assaf and Turco, 1994). The authors concluded that AT mineralization was controlled by soil C levels. The concentration of metabolites found was also related to the amount of soil C. The results of this study suggest a concomitant metabolism of AT and AL by the SMC. The C and N requirements of the microorganisms results in AT dependent AL degradation. Another indication of the inter-relationship between AT and AL biodegradation is the lack of linearity of the semi-logarithmic degradation plots (Figs. 2B and 3B). Typically pseudo-first order degradation, which is usually associated with cometabolism, results in linear relationship when degradation is represented as the natural log of the fraction of substrate present [LN (C/C0)] over time (Shelton and Doherty, 1997). Degradation of both AT and AL slowed considerably reaching a plateau at about 2 weeks for AT and at 4 weeks for AL. There are several possibilities for this diminished degradation. The lack of other nutrients or growth regulators may cause the microbes to become inactive. Addition of soil or yeast extract to the minimal medium used to enumerate pesticide degraders increased the number of active organisms (Fourner et al., 1997). Another reason for the plateau might be related to the nature of the AT and AL contamination. Usually, the amount available to microbes decreases with time until a minimum value is reached, at which point no more degradation of the containment will occur. Once the microorganisms degrade the available substrate present in the soil, the degradation rate is limited to the rate of desorption of the sorbed fraction back into the soil solution (Alexander, 2000). 4.2. Metabolite production and degradation All of the compounds tested for except deethylatrazine (CIAT) were detected during incubation (Fig. 4a; Table 4). The high concentration of hydroxydeethylatrazine (OIAT) might have indicated that CIAT might also have been transformed to OIAT. The presence of both deethyldeisopropylatrazine (CAAT) and deethylhydroxyatrazine (OIAT) suggested that CIAT was formed but was a transient metabolite, which did not buildup in the soil. It

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Table 4 AT metabolite production and degradation over time during the SMC degradation study. ‘‘C formed’’ is the amount of metabolite produced while ‘‘C deg’’ is the amount of the metabolite that is further degraded. The ‘‘DC’’ represents the actual concentration of the metabolite; i.e., the mM formed minus the mM degraded. The two percentages are based on 1) the total amount of AT degraded throughout the study (% ATd) and 2) the amount of AT initially present in the soil (% AT0). Day

OIET

0 3 5 7 14 21 28 37 118 133

0 0 0.0443 0 0 0.2805

Day

OAAT

0 3 5 7 14 21 28 37 118 133

0 0 0.0225 0.0000 0.0000 0.0262

C formed

0.0443

0.0938 0.1983

DC

% ATd

% AT0

C formed

0 0 0.0443 0 0 0.2805 0.2005 0.2005 0.2859 0.0876

0 0 11.64 11.64 11.64 73.70 52.68 50.47 75.12 23.02

0 0 5.60 5.60 5.60 35.47 25.35 24.28 36.15 11.08

0 0.0223 0.0838 0 0.1030 0.0305 0 0 0.0772

DC

% ATd

% AT0

0.0772

0 0.0223 0.1061 0 0.1030 0.1335 0.0795 0 0.0722 0

0 5.86 27.88 0 27.06 35.08 20.89 0 20.28 0

0 2.82 13.42 0 13.02 16.88 10.05 0 9.76 0

C Deg.

DC

% ATd

% AT0

0 0 0.0064 0 0 0.0114 0.0349 0.0210 0.0119 0.0406

0 0 1.68 0.00 0.00 3.00 9.17 5.52 3.13 10.67

0 0 0.81 0 0 1.44 4.41 2.66 1.50 5.13

C Deg.

0.1061

0.0540 0.0795

CEAT C Deg. 0 0.0315 0.0160 0.0164

0.0288 0

0.0119 0.0159

DC

% ATd

% AT0

0 0 0.0225 0 0 0.0262 0.0098 0.0386 0.0267 0.0108

0 0 5.91 0 0 6.88 2.57 10.14 7.02 2.84

0 0 2.84 0 0 3.31 1.24 4.88 3.38 1.37

C Formed 0 0 0.0064 0 0 0.0114 0.0235 0 0 0.0287

0 0.0202 0.0104 0 0.0139 0.0091 0

CAAT C Formed

0 3 5 7 14 21 28 37 118 133

C Deg.

0.0800 0.0084

C Formed

Day

OIAT

0 0 0.0224 0 0.0015 0.0041 0.0040 0.0019 0 0.0305

C Deg. 0 0.0227 0 0 0 0.0045 0

DC

% ATd

% AT0

0 0 0.0224 0 0.0015 0.0056 0.0096 0.0115 0.0070 0.0375

0 0 5.89 0 0.39 1.47 2.52 3.02 1.84 9.85

0 0 2.83 0 0.19 0.71 1.21 1.45 0.89 4.74

seemed that two possible degradation pathways were present in the SMC culture. One pathway involved dealkylation and deamination as the first step (i.e., the presence of CEAT and CAAT) while the other pathway had the hydrolytic dehalogenation of atrazine to hydroxyatrazine (OIET) [i.e., the presence of OIAT, OAAT and the high amount of OIET]. By one week, all of these metabolites disappeared from the biometer flasks. Because deamination rapidly follows dealkylation, the combined process of dealkylation/deamination was considered a major step in the AT degradation pathway. Therefore, it was not unexpected to see all of the detected metabolites disappear by Day 7. All of the hydrozylated metabolites (OAAT and OIAT) formed were subsequently transformed, while 78% of deethylatrazine (CEAT) and 54% of the deethyldeisopropylatrazine (CAAT) was further degraded. This indicated that deamination of both the hydroxylated and non-hydroxylated metabolites occurred. The hydroxylated metabolites were transformed at a greater rate resulting in zero accumulation. Deisopropylatrazine (CEAT) is a metabolite in the breakdown of both cyanazine (CY) and simazine (SI) as well as AT. Both of these degradation processes contributed to the appearance of CEAT, which might be further transformed to deethyldeisopropylatrazine (CAAT). Some bacteria prefer CY over AT. Gebendinger and Radosevich (1999) found that degradation of AT by Ralstonia spp. M91-3 was competitively inhibited by the presence of CY while SI had no effect on AT degradation. The buildup of hydroxyatrazine (OIET) at the end of the incubation period occurring at the same time as the

increase in both CEAT and CAAT might have been indicative of the inhibitory effect of CY on the further degradation of OIET. The microbes might have been selectively degrading CY at the expense of OIET. Prior exposure to OIET inhibited the AT degrading activity of washed cell suspensions of the M91-3 isolate, which is able to utilize AT as the sole source of C and N (Gebendinger and Radosevich, 1999). Grigg (1997) monitored AT degradation in the presence of other compounds typically found at mix-load sites. Single co-contaminants had no effect on AT degradation rates. However, mixtures of 2 co-contaminants increased the half-life of AT and hydroxyatrazine (OEIT), compared to when AT was the single contaminant. Atrazine degradation in the presence of AL and ME increased t1/2 from 1.0 d to 1.4 d. In more complex mixtures, where trifluralin was mixed with either AL or ME, the AT t1/2 increased to 2.3 d and 2.0 d, respectively. There was a 2–3 fold increase in t1/2 when all three herbicides were present. The presence of the other contaminants in the mix-load site soil used in these experiments must certainly have had an effect on AT degradation considering that AL, ME, CY and SI were all degraded. It is possible that the presence of these cocontaminants reduced the amount of AT completely mineralized because more easily degraded substrates were available to the microbes (side chains of the other s-triazines, for example). But the SMC developed for AT and AL degradation was able to adapt and use many other compounds as substrates. This was a positive attribute that can be advantages during in-situ remediation efforts.

A.E.M. Chirnside et al. / Soil Biology & Biochemistry 41 (2009) 2484–2492 Table 5 AL metabolite production and degradation over time during the SMC degradation study. ‘‘C formed’’ is the amount of metabolite produced while ‘‘C deg’’ is the amount of the metabolite that is further degraded. The ‘‘DC’’ represents the actual concentration of the metabolite; i.e., the mM formed minus the mM degraded. The two percentages are based on 1) the total amount of AL degraded throughout the study (% ALd) and 2) the amount of AL initially present in the soil (% AL0). DAY

DMA

0 3 5 7 14 21 28 37 118 133

0 0.014 0.046 0 0 0.046

DAY

DIE

0 3 5 7 14 21 28 37 118 133

0 0 0.019 0.080 0 0.017 0.052 0.076 0.007 0

C Formed

C Deg.

0.038 0.007 0.014

0.071 0 0.030

C Formed

0.084 0

DC

% ALd

% AL0

0 0.014 0.060 0.021 0.015 0.061 0.047 0.118 0.034 0.064

0.00 2.44 10.71 3.85 2.70 10.91 8.48 21.13 6.13 11.43

0.00 1.80 7.89 2.83 1.99 8.04 6.25 15.58 4.52 8.42

C Deg.

DC

% ALd

% AL0

0 0.014 0 0

0 0 0.019 0.099 0.085 0.102 0.154 0.229 0.236 0.150

0.00 0.00 3.40 17.72 15.22 18.35 27.60 41.19 42.46 27.03

0.00 0.00 2.50 13.06 11.22 13.52 20.34 30.36 31.30 19.92

0 0.086

Both of the two AL metabolites were detected throughout the incubation period (Fig. 4b, Table 5). Initially more 2-chloro-20 ,60 diethylacetamilide (DMA) was seen in the biometer soil than 2,6diethylanaline (DIE). However, the DMA concentration decreased from Day 5 to Day 14 while DIE increased in concentration. The formation of DMA has been considered a minor AL degradative pathway with subsequent degradation to DIE or 1-chloroacetyl-1,2dihydro-7-ethylindole proceeding at a slow rate. In this incubation, DMA was transformed quickly with an accompanying increase in DIE suggesting DMA is transformed to DIE. Diethyl aniline (DIE) is considered the major metabolite of AL degradation with 2 degradative pathways leading to its formation (Guo and Wagenet, 1999). In this study, 70% of the AL was rapidly transformed (t1/2 ¼ 8 d, Fig. 4b) suggesting that the formation of DIE from DMA was the major degradation pathway. Metolachlor was degraded at a much slower rate than the AL in the contaminated soil (Fig. 5b). It has been shown in the literature that ME is held more tightly to the soil than AL. 5. Conclusions Inoculation of the pesticide-contaminated mix-load site soil with the isolated indigenous SMC resulted in the degradation of both AT and AL. Approximately 48% of the AT (0.381 mM) and 70% of the AL (0.527 mM) were transformed within the biometer flasks during the 133 d incubation period. The net CO2 evolved (0.948 mM) represented 10.3% of the total CO2 evolved if AT and AL were completely mineralized. The concentration of SMC organisms remained steady with no significant increase in population during incubation indicating cometabolic biodegradation. The calculated pseudo first-order rate constant was 0.00064 d1 and 0.0814 d1 for AT and AL, respectively. Initially AL decreased at a slower rate than AT and its degradation appeared to be dependent upon AT degradation. The calculated half-life for AT was 108 d.

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Alachlor degradation exhibited a much shorter half-life of approximately 9 d. Two possible degradation pathways for AT were evident from the metabolite concentrations detected. One pathway involved dealkylation and deamination as the first step while the other pathway began with the hydrolytic dehalogenation of AT to hydroxyatrazine (OIET). Both of the two AL metabolites (demethoxymethyl-alachlor, DMA and 2,6-diethyl aniline, DIE) were detected throughout the incubation periods. The DMA was rapidly transformed to DIE. Thus DIE was one of the major metabolites of AL degradation. The SMC was also able to degrade 16%, 55% and 20% of the initial cyanazine (CY), simazine (SI), and metolachlor (ME) found in the contaminated soil, respectively.

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