Kinetics of biodegradation of nitrilotriacetic acid (NTA) in an estuarine environment

ECOTOXICOLOGY

AND

ENVIRONMENTAL

SAFETY

12,166- 179 ( 1986)

Kinetics of Biodegradation of Nitrilotriacetic in an Estuarine Environment ROBERT J. LARSON*

Acid (NTA)

AND ROY M. VENTULLO~

*Environmental Safety Department, Procter & Gamble Company, Ivorydale Technical Center, Cincinnati, Ohio 45217, and TDepartment ofBiology, University ofDayton, Dayton, Ohio 45469 Received April 17, 1986 The effects of salinity and dissolved organiccarbon (DOC) on the kinetics of biodegradation of nitrilotriacetic acid (NTA) were studied in a Canadian estuary with a prior history of NTA exposure. Kinetic parameters for degradation of L4C-labeled NTA, maximum velocity (V-) and first-order rate constant (k,), were estimated by nonlinear regression models from velocity and time-course plots, respectively. The distribution of bacteria with NTAdegrading capability was also determined at various salinities and DOC levels by the “C-most-probable-number (14CMPN) technique. In general, NTA degradation was rapid in estuarine water over the range of salinities and DOC levels tested. Mean V,, and k, values (+ standard deviation) across several sampling periods averaged 4753 + 2849 ng liter-’ hr-’ and 0.32 + 0.19 day-‘, respectively. The estimated half-life for NTA degradation in estuarine water, based on the mean kl value, was -2 days. Degradation rates for NTA were relatively insensitive to changes in salinity or LKK values, and neither of these two parameters had signilicant effects on NTA degradation at the microbial community or individual cell levels. Based on 14C-MPN results, the distribution of estuarine bacteria capable of degrading NTA was broad and not related to salinity or DOC levels. The NTA degraders appeared to be indigenous members of the estuarine microbial community and not wastewater-associated microorganisms. 0 1986 Academic Ress, Inc.

INTRODUCTION The biodegradability of metal complexes of nitrilotriacetic acid (NTA), an amino polycarboxylic acid used as a builder in synthetic laundry detergents, has been studied by a number of investigators for almost 20 years. In general, the results of these studies indicate that NTA-metal chelates are biodegraded in a variety of environmental systems, including wastewater treatment systems (Forsberg and Lindquist, 1967; Swisher et al., 1967, 1973; Bouveng et al., 1968, 1970; Thompson and Duthie, 1968; Shumate et al., 1970; Hubly and Cleasby, 197 1; Bjorndal et al., 1972; Eden et al., 1972; Huber and Popp, 1972; Rudd and Hamilson, 1972; Cleasby et al., 1974; Klein, 1970; Renn, 1974; Gudematsch, 1975; Shannon, 1975; Walker, 1975; Shannon et al., 1978; Stoveland et al., 1979; Wei et al., 1979; Obeng et al., 1982; Vashon et al., 1982; Stephenson et al., 1983), soils (Tiedje and Mason, 1974; Tabatabai and Bremner, 1975; Means et al., 1980; Kirk et al., 1983b), and surface waters (Chau and Shiomi, 1972; Warren and Malec, 1972; Shannon et al., 1974; Bott et al., 1980; Larson et al., 198 1; Larson and Davidson, 1982). These studies and others have shown that degradation of NTA occurs over a range of conditions of pH, temperature, water hardness, dissolved oxygen concentrations, and NTA-metal concentrations (Claesson, 197 1; Wong et al., 1972; Enfors and Molin, 1973a,b; Liu et al., 1973; Moore and Barth, 1976; Larson et al., 198 1). Degradation rates exhibit kinetic and thermodynamic patterns which are typical of those observed for natural organic compounds (Thompson and Duthie, 1968; Shannon et al., 1978; Larson et al., 198 1). Studies by 0147-6513186 $3.00 Copyright Q I986 by Academic Press, inc. All rights of reproduction in any form reserved.

166

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several investigators have also shown that NTA is biodegraded by indigenous microorganisms present in groundwater aquifers and in subsurface soil systems (Dunlap et al., 1972; Hrubec and Van Delhi, 198 1; Larson and Ventullo, 1983; Ventullo and Larson, 1985; Ward, 1985). Degradation of NTA in subsurface environments occurs at relatively rapid rates, given the reduced microbial activity in these systems in general, and is observed under both aerobic and anaerobic conditions. In total, microbial biodegradation has been shown to be an effective removal mechanism for NTA in essentially all environmental compartments exposed to this material via its use in and disposal of detergent products. The only major environmental compartments where questions still remain about the biodegradability of NTA include mid-salinity coastal marine and estuarine waters. Contradictory results have been reported in the literature regarding the biodegradability of NTA in saline environments. Bourquin and Przybyszewski (1977) were unable to isolate bacteria from a gulf coast estuary (Escambia Bay, Fla.) capable of metabolizing NTA as a sole carbon and energy source. They were, however, able to isolate NTAdegrading bacteria from adjacent freshwater environments and suggested that the lack of NTA degradation in estuarine water was due to inhibition of NTA-specific enzymes by high salt, i.e., ~10 ppt. Kirk et al. (1983a) also failed to demonstrate NTA degradation in laboratory simulation studies where bacteria of marine origin were exposed to NTA in a synthetic sewage/seawater matrix. Erickson et al. ( 1970), however, were able to isolate bacteria from seawater which were capable of growth on NTA as a sole nitrogen source or nitrogen and carbon source. Estuarine bacteria capable of metabolizing NTA have also been recently identified by microautoradiographic techniques in coastal salt marshes in the southeastern United States (Pfaender and Paerl, 1984). Due to the different experimental conditions used in many prior NTA biodegradation studies, it is difficult to state with certainty the most important factor controlling the rate and occurrence of NTA degradation in different saline environments. However, one factor which has already been shown to play a key role in freshwater environments, and is probably also important in saline environments, is microbial acclimation. Numerous studies in wastewater treatment systems, soils, and surface waters have shown that acclimation of microbial populations to NTA is necessary before rapid degradation can occur. The length of this acclimation period is variable, ranging from a few days to several weeks, and appears to involve both microbial population growth and enzyme induction or recruitment (International Joint Commission, 1978). Recent studies by Pfaender et al. (1985) have shown that acclimation can be an important factor controlling the kinetics of NTA degradation in some estuarine environments. Microbial populations in estuarine water not previously exposed to NTA required 5-7 weeks to develop maximum NTA-degrading capability. By contrast, populations in estuarine waters naturally exposed to NTA in wastewater effluent exhibited an immediate degradation response and no acclimation period was necessary. In this paper, we report the results of studies to characterize the kinetics of NTA biodegradation in an estuarine ecosystem with a prior history of NTA exposure. The studies were conducted over a 14-month period and were designed to measure the rate and extent of NTA degrada;tion over a range of conditions of salinity, dissolved organic carbon, and microbial biomass levels. The specific objectives of the studies were threefold:

168

LARSON AND VENTULLO /d

- Zone of major influence of sewage effluent

0

- Study Area

VANCOUVER

LULU

I

ISLAND

TSAWWASSEr;’ **- \ FERRY TERMINAL

FIG.

1. Fraser River Estuary and location of study site.

(1) The determine the distribution of NTA-degrading bacteria as a function of salinity in an estuarine system exposed to NTA-containing wastewater. (2) To characterize the heterotrophic activity of estuarine bacteria on NTA and the effects of salinity and nutrient (DOC) levels on NTA metabolism at the microbial community and individual cell levels. (3) To characterize the kinetics of NTA degradation in estuarine water over a range of salinities and DOC levels at NTA concentrations approximating realistic environmental levels. MATERIALS

AND

METHODS

Study site and sample collection. Estuarine water samples for all biodegradability studies were collected from the Fraser River Estuary (FRE), Sturgeon Bank, near Vancouver, British Columbia (Fig. 1). The Fraser River Estuary is a relatively shallow (3- 10 m), wind-driven estuary which receives freshwater input from the Fraser River and wastewater effluent from the Iona Island Municipal Wastewater Treatment Plant

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ACID BIODEGRADATION

169

(WTP). The Iona Island WTP is a medium-sized primary treatment facility with an average flow of about 100 million gal day -‘. Water samples used in biodegradability studies were collected at depths of l-2 m, approximately 100-300 m offshore at various distances (300-4600 m) from the outfall of the Iona Island Wastewater Treatment Plant. The temperature (“C) was recorded at the time of collection and salinity (ppt) was measured on-site with a YSI Model 33 salinometer. Water samples were shipped by air to Cincinnati and stored at 4°C prior to testing, which began 8- 12 hr after sample collection. Chemicals. The nitrilotri-[U-‘4C]acetic acid, trisodium salt (14C-NTA) used in testing was obtained from Amersham, Inc. (Arlington, Va.) with a specific activity of 42 &i mg-’ (as the monohydrate). It was synthesized by reacting chloroacetic acid with [U-‘4C]glycine in the presence of NaOH and was purified by crystallization. As received, the radiochemical purities of this material by reverse isotope dilution and TLC were 95 and 98%, respectively. Reanalysis of this material in our laboratories by reverse-phase ion pair liquid chromatography (Parks et al., 198 1) indicated a radiochemical purity of 95%. The remaining 5% radiolabeled impurity was identified [14C]glycine. D-[U-14C]Glucose was obtained from Amersham with a radiochemical purity of >99% and a specific activity of 1600 &i/mg. All other chemicals used in testing were reagent-grade materials, obtained from commercial sources. Analytical procedures. Dissolved organic carbon (DOC) was measured in purged, acidified (
(1) where f = (14C uptake + “C02)/‘4C added, t = incubation time (hours), A = added substrate concentration (&liter), I’,, = maximum degradation rate (ng/liter/hr), and K, = half saturation constant (pg liter-‘) where u = 0.5 V,, . Parameter estimates

170

LARSON

AND VENTULLO TABLE

CHARACTERIZATION

1

OF FRASER RIVER ESTUARY WATER USED

IN BIODEGRADATION STUDIES Parameter Temperature PB Salinity

Range of values

unit

15-21 7.0-8.5 4-19

‘C mg 100 ml-’ mg liter-’ hr

1.7-11.5

Glucose Tn AODC MPN-NTA

1.8-8.4 degraders

2.8 x lo’-1.3 3.2 x lOI-7.8

X 10’ x 10’

cells ml-’ cells ml-’

for V, and K, were obtained by least-squares analysis using a nonlinear computer program, as previously described (Larson, 1984). Estimated V,, values were divided by the number of total bacteria (AODC) or MPN of NTA degraders to generate V,, specific activity indices (Wright, 1978). Time-course studies were conducted as previously described (Larson and Payne, 198 1). Briefly, i4C-NTA (N 10 to 100 pg liter-‘) was added to triplicate l- (or 2-liter Erlenmeyer flasks containing 0.5 or 1 liter of estuary water, respectively. Flasks were incubated at 19 f 2°C in a constant-temperature room (Forma Scientific, Cleveland, Ohio) and constantly agitated (150 rpm) on a rotary platform shaker. At various intervals, subsamples were taken from solution and the amount of radioactivity was determined in three fractions: 14COz, 14C activity in biomass, and 14Cactivity remaining in solution. Mass balances were made at each sampling point and ranged from 80 to 99%. Data for cumulative percentage 14C02 produced or 14C activity removed from solution were analyzed by nonlinear regression models to estimate the tlrstorder rate constant &) for mineralization and removal, normalized for the extent of biodegradation observed. Parameter estimates for k, were generated by iterative techniques using least-squares analysis and a nonlinear computer program, as pre viously described (Larson, 1984). H&lives (t& were determined by the relationship t L,z= In 2/k,. RESULTS

Fraser River Estuary Site Characterization data for the estuarine water used in NTA biodegradability studies are summarized in Table 1. In general, the range of values observed for FRE water is typical of an exposed coastal estuary receiving significant quantities of freshwater input as well as wastewater discharges. The salinity varied ---fold (4-19 ppt) and total organic carbon (TOC) levels varied from -2 to 12 m&liter. Glucose turnover times, a measure of general bacterial heterotrophic activity, were relatively rapid (2 to 8 hr), indicating a high level of metabolic activity in the water samples tested. Total cell numbers, based on AODC, varied -4O-fold; the MPN of NTAdegrading bacteria showed less variation than did AODC, or -25-fold. There was no apparent relationship between the number of NTA degraders and salinity or the number of total cells by AODC at different locations in the estuary (Table 2). High MPN values were obtained at both low and high salinities, and the ratio of NTA MPN values to

ESTUARINE

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ACID

171

BIODEGRADATION

TABLE 2 NUMBER

OF TOTAL

AND NTA-DEGRADING

BACTERIA

MPN-NTA

DiStanCe

Date September

from

WTP

1983

June 1984

IN FRASER

(m)

salinity

(x10*

300 400 1200 1900

8 11 15 19

300 600 1600 2600

4 8

RIVER

ESTUARY

degraders’

AODCb (X106cellsml-‘)

cells ml-‘)

0.3 2.5 0.8 7.8

WATER

(0.1-0.9) (0.9-6.7) (0.3-2.4) (2.5-24.3)

12.6 (12.0-13.2) 6.5 (3.9-9-l) 0.5 (0.3-0.8) 0.3 (0.1-0.4)

5.2 (2.0-13.5) 3.3 (1.3-8.6) 1.4 (0.5-4.3) 0.9 (0.3-3.0)

9.2 (4.4-14.0) 3.8 (1.2-6.4) 4.7 (2.9-6.5) 3.0(1.5-4.5)

(1Most-probable-number of NTA-degrading bacteria; values in parentheses are 95% confidence intervals of mean values. b Acridine orange direct counts; values in parentheses are 95% confidence intervals of mean values.

AODC varied significantly between different sites and sampling periods. The distribution of NTAdegrading bacteria in FRJZ water was broad, therefore, and not related to the distance f%om the point of wastewater input.

Eflects of Salinity and DOC on NTA Heterotrophic Activity Kinetic parameters for biodegradation of NTA in short-term heterotrophic activity assays are summarized in Table 3. A typical plot showing the initial rate of NTA degradation as a function of concentration is given in Fig. 2. For all water samples tested, the rate of NTA biodegradation (14C uptake + 14C02 production) was a saturable function of NTA concentration. Rate data were accurately described by Eq. [2], and correlation coefficients for the fit of experimental data to the predicted curves were 0.97 or greater (Table 3). T’he maximum rate of NTA degradation was high and TABLE 3 HETEROTROPHIC

Acrrvrn

OF E~TIJARINE BACTERIA AND DOC LEVELS

ON NTA

AT VARIOUS

Specific Distance

Salinity DOC (ppt) (mg liter-‘)

Date

from WTP (m)

1984

300 600 1600 2000

4 8 t

11.5 4.0 2.2 1.7

1200 1800 2200 3600 4600

6 9 12 15 17

6.4 8.6 3.2 6.4 6.6

June

August

1984

.. Klllu (ng liter-’ hr-‘)

activityindex

MPN (10”

fig cell-’

6393(1584) 1356 (709) 1049 (122) 602 (58)

12.3 4.1 7.5 6.7

5785 (2086) 6714121 IO) 7073 (2223) 7201(2213) 6607 (2233)

-b -

SALINITIES

AODC hr-‘)

(IO-‘*

pg cell-’ 695 357 223 7.01 1808 2098 1912 2400 2753

hr-‘)

r’ 1.00 0.99 0.99 0.99 0.97 0.98 0.98 0.99 0.98

172

LARSON

AND VENTULLO

NTA CONCENTRATION

(Ml”)xlO’

2. Kinetics of NTA degradation in heterotrophic activity studies. Data are for water collected in June 1984 at the 2600-m site, and parameter estimates for V,, and K,, are given in inset. Dotted contours are 95% confidence limits of true mean. FIG.

relatively constant for all waters tested, averaging 4753 + 2849 ng liter-’ hr-’ across all sampling sites and periods. There was no consistent effect of salinity or DOC levels on NTA degradation rates. The V,, at low (4 ppt) and high (17 ppt) salinities was virtually identical in different water samples, and within a given sampling period (i.e., 8/84) the variation of V,, with salinity and DOC was relatively small. A significant decrease in V,, was apparent at one sampling period (6/84) and this decrease appeared to correlate well with decreasing DOC values and distance from the WTP outfall. However, the MPN of NTAdegrading bacteria also decreased with DOC levels in these water samples (Table 2), and the decrease in MPN was presumably responsible for the observed reduction in V,, values. The maximum rate of NTA degradation on a per cell basis, based on I’,,,, specific activity indices (SAI), varied a maximum of about IO-fold during different sampling periods (Table 3). However, within a given sampling period, SAI values showed much less variation, ---fold in the 6/84 sample and 12-35s in the 8/84 sample. The variation in SAI was about the same irrespective of whether AODC or the MPN of NTA degraders was used to calculate activity indices. Both enumeration techniques, therefore, provided comparable information concerning the activity of microbial communities on a per cell basis. Based on AODC, V,, SAI values for NTA were relatively insensitive to changes in salinity or DOC values and neither parameter had consistent effects on SAI (Table 3). For example, SAI appeared to increase with decreasing salinity in the 6/84 sample, khereas it decreased with decreasing salinity in the 8/84 sample. The SAI increased with increasing DOC in the 6/84 sample, but was virtually constant at the extremes of DOC measured in the 8/84 sample. In general, no consistent adverse effects of salinity or DOC were observed on V,, or SAI values over the range of conditions tested. Biodegradation Kinetics-Time-Course Studies The distribution of 14C activity during biodegradation (2600-m site) at an initial concentration of 11 pg liter-’

of 14C-NTA in FRE water is given in Fig. 3. Data for

ESTUARINE

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ACID BIODEGRADATION

173

TIME (DAYS) FIG. 3. Kinetics of NTA degradation in time-course studies at an iniial concentration of 11 pg liter-‘. Data are for water collected in September 1983 at the 2600-m site, and the parameter estimate for k, is given in Table 4. Dotted contours are 95% confidence limits.

the rate and extent of NTA mineralization and removal over a lo-fold range of NTA concentrations are summarized in Table 4. In general, degradation of 14C-NTA followed apparent first-order kinetics over the NTA concentration range tested. Kinetic patterns for removal of 14Cactivity from solution were comparable to those for 14C02 evolution, and rate constants for removal and mineralization were not significantly different (P < 0.01). Over the range of NTA concentrations tested, degradation occurred immediately with no apparent lag phase. This indicates that FRE microbial communities were adapted to NTA, presumably as a result of prior exposure to NTAcontaining wastewater. During biodegradation, high levels of 14C02 were produced (-90%) and relatively little 14C label was incorporated into biomass (- 10%). This indicates that FRE microbial communities were primarily using the NTA carbon skeleton for energy production in dissimilatory metabolic pathways. Eflects of Salinity Table 4 also summarizes data for the rate and extent of NTA biodegradation across all sampling periods as a function of salinity. In springwater samples (6/84), large amounts of freshwater input from the Fraser River maintained the salinity at low and relatively constant values at all sampling sites. Degradation rates for NTA showed relatively little variation at different locations, with a mean first-order rate constant + standard deviation (SD) of 0.42 + 0.17 day-‘. In the summer water samples (Sep tember 1983 and August 1984), a larger range of salinities was observed at different sites due to reduced freshwater input. However, within a given sampling period (i.e., June 1984) degradation rates for NTA were essentially identical at both high and low salinities. Across all sampling sites and salinities tested, the extent of 14C02 production (*SD) averaged 88.3 f 5.0% and the mean first-order rate constant for mineralization (*SD) averaged 0.32 f 0.19 day-‘. The estimated half-life for NTA in estuarine waters, based on this mean kr value, was -2 days.

Length of

21 28

6 17

8 8

4 8

19 19

(PPt)

Salinity

13 12

11 11

11 11

11 108

NTA concn (pgliter-‘)

0Values in parentheses represent 1 standard deviation. bValues in parentheses represent thestandard errorsof theparameter estimates.

1800 4600

18 18

1600 2600

August 1984

18 18

300 600

June 1984

30 30

2600 2600

experiment WYS)

September1983

Date

Distance from WTP (m)

KINETIC PARAMETERS FOR BIODEGRADATION

TABLE

87.3 (1.5) 78.1 (5.0)

89.2(1.8) 91.1 (2.4)

90.2 (1.9) 92.6 (1.2)

93.7 (1.2) 94.6 (0.6)

STUDIES

0.12 (0.06) 0.11 (0.05)

0.23 (0.02) 0.33 (0.03)

0.64 (0.1) 0.47 (0.04)

0.2 1(0.03) 0.44 (0.11)

Rate constant (days-‘)b

OF NTA IN TIME-COURSE

14c02 produceda (% of 14Cadded)

4

5.6 (0.4) 9.9(1.3)

5.5 (0.8) 4.9 (1.4)

4.3 (0.2) 3.5 (0.5)

3.6(1.0) 2.4 (0.1)

14Cin solution’ (% of 14Cadded)

0.10 (0.08) 0.08 (0.07)

0.15 (0.04) 0.30 (0.02)

0.57(0.11) 0.40 (0.06)

0.18 (0.02) 0.43 (0.10)

Rateconstant (days-‘)b

”

2 F:

u s

$ $

i

ESTUARINE

NITRILOTRIACETIC

ACID BIODEGRADATION

175

DISCUSSION Previous studies to determine the biodegradability of NTA in saline environments have yielded mixed results. Although NTA degradation has been shown to occur in high-salinity (35 ppt) marine systems (International Joint Commission, 1978), the metabolism of NTA by mid-salinity estuarine microbial communities has not been clearly demonstrated. Erickson et al. (1970) reported that bacterial populations in natural seawater were capable of metabolizing NTA as a sole carbon and nitrogen or a sole nitrogen source. Using microautoradiographic techniques, Pfaender and Paerl (1984) also identified estuarine bacteria capable of metabolizing NTA. However, other investigators were unable to isolate bacteria capable of metabolizing NTA from artificial seawater or estuary water. They postulated that the lack of NTA degradation was due to the absence of NTA degraders in the marine bacterial population (Kirk et al., 1983a) or the inhibition of enzymes specific for NTA by high ionic strength (Bourquin and Pxzybyszewski, 1977). Another possibility which was not considered in these studies, but which is a factor in all laboratory investigations, is that the experimental conditions were not sufficient to develop acclimated populations of NTAdegrading bacteria. Recent studies by Pfaender et al. (1985) showed that estuarine bacteria previously unexposed to NTA may require extended periods to adapt to this material in the laboratory. They also found that the relative rates of adaptation to NTA did not correlate well with any ofthe physical/chemical or general microbiological characteristics of the estuaries that were measured. To avoid laboratory acclimation effects, we chose to study an estuarine environment with a prior history of NTA exposure. Our results clearly indicate that estuarine waters exposed to NTA in wastewater effluent have significant concentrations of NTA-degrading bacteria and a high level of degradation activity. The NTA-degradation activity is broadly distributed throughout the estuary and is not related to the distance from the point of wastewater input. The NTA-degrading bacteria, therefore, are not wastewater-associated microorganisms, but appear to be part of the indigenous estuarine microflora which have developed as a result of exposure to NTA in wastewater discharges. Two different approaches were used to measure the kinetics of NTA degradation in estuarine water. The V,, , or heterotrophic activity, approach was used to estimate maximum biodegradation rates at saturating NTA concentrations where the kinetics of degradation were zero-order and therefore independent of NTA concentration. Relatively short incubation times were used to avoid alterations in the microbial community and V,, as a result of confinement effects. The V,,, was then divided by number of total bacteria (AODC) to estimate SAI values. Wright has proposed that SAI values can serve as an indicator of the “physiological state” of individual bacteria, allowing the effects on metabolic activity to be determined on a per cell basis. In our studies, we used SAI to examine the effects of salinity and DOC on rates of NTA degradation at the level of individual cells. Since V,, is an extrapolated value measured at exaggerated substrate concentrations which may not in practice be attained in the environment, a second kinetic parameter, the biodegradation rate constant (k,), also was used to estimate rates of NTA biodegradation. Values fork, were measured at low subsaturating NTA concentrations, where the kinetics of degradation were first order and therefore directly proportional to NTA concentration. At low NTA concentrations, which approximate

176

LARSON

AND VENTULLO

realistic environmental levels, half-lives for NTA degradation can be directly calculated from the corresponding first-order biodegradation rate constant. In general, V,, values for NTA degradation in the Fraser River Estuary were significantly higher (3- to 40-fold) than previously reported values for a coastal estuary (Newport River Estuary) in the southeastern United States (Pfaender and Bartholomew, 1982; Bartholomew and Pfaender, 1983). The U.S. estuary had similar physical/chemical and general microbiological (AODC, Tn) characteristics to the Canadian estuary, but had no known history of NTA exposure. This was presumably responsible for the observed differences. The V,, and SAI for NTA degradation in FRE water were relatively insensitive to changes in salinity or DOC within a specific sampling period, and neither parameter had a significant adverse effect on NTA degradation at the microbial community or individual cell level. The Vmax, however, did tend to decrease as the MPN of NTA degraders decreased. This suggests that the number of potential NTA degraders is more important than factors like salinity or DOC in controlling NTA degradation rates. These results are similar to those reported by Pfaender et al. (1985), who studied adaptation of microbial communities to NTA in several coastal estuaries. They found that although adaptation to NTA could occur in estuarine water samples where essentially no NTA degraders could be detected by the MPN technique, the only variable tested which directly correlated with adaptation times was the MPN of NTA degraders. Rates of NTA degradation in time-course experiments were relatively constant and followed apparent first-order kinetics over the range of NTA concentrations tested. The mean half-life for degradation in FRE water ( - 50 hr) was somewhat longer than values recorded in well-acclimated river water (- 10 hr) with a comparable history of NTA exposure (Larson and Davidson, 1982). This result is consistent with earlier reports by Pfaender and Bartholomew (1982), which indicate that estuarine microbial communities were somewhat less active against NTA than were freshwater communities. As observed for V,, , the observed first-order rate constants for NTA degradation were relatively insensitive to salinity, DOC, and AODC levels, both within and between sampling periods (Table 4). Degradation occurred over a range of salinities and biomass levels, and at DOC levels as high as - 12 mg/liter. The latter result indicates that NTA degradation occurs concurrently with DOC metabolism, even at initial DOC concentrations that are 1OO-to lOOO-fold higher than initial NTA concentrations. Competitive inhibition of NTA degradation by other nutrients, therefore, is not likely to occur even at very high nutrient to NTA ratios. SUMMARY

AND

CONCLUSIONS

Previous studies have suggested that degradation of NTA does not occur in estuarine environments due to a lack of potential NTA-degrading bacteria or to the inhibition of the enzymes involved in NTA degradation by salt. A third possibility, which was not addressed by previous studies, is that suitable conditions were not provided in the laboratory for the development of adapted populations of NTAdegrading bacteria. To avoid laboratory acclimation effects, we studied the kinetics of NTA biodegradation in a Canadian coastal estuary with a prior history of NTA exposure. Based on the results of our studies, the following conclusions can be drawn:

ESTUARINE

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ACID BIODEGRADATION

177

( 1) The number (MPN) of bacteria capable of degrading NTA was significant ( 1021O3cells ml-‘) and broadly distributed throughout the estuary. There was no apparent relationship between the MPN of NTA degraders and salinity, DOC, or AODC levels, or the distance from the point of wastewater input. (2) The heterotrophic activity of estuarine bacteria on NTA was high and relatively constant across all waters tested. No adverse effects of salinity or DOC levels were observed on heterotrophic activity at either the community (I’,,) or individual ceil (SAI) levels over the range of conditions tested. (3) The rate and extent of NTA degradation in estuarine water were relatively insensitive to differences in salinity, DOC, or AODC levels. Degradation followed apparent first-order kinetics over the NTA concentration range tested, with a mean halflife for degradation across all waters of - 2 days. ACKNOWLEDGMENTS We thank Pamela &scone, Daniel Davidson, and Theresa John for their help in completing the technical aspects of this work.

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