groundwater infiltration: Laboratory column studies

War. Res. Vol. 21, No. 10, pp. 1237-1248, 1987 Printed in Great Britain. All rights reserved

0043-1354/87 $3.00+0.00 Copyright © 1987PergamonJournals Ltd

MICROBIAL DEGRADATION OF NITRILOTRIACETATE (NTA) DURING RIVER WATER/GROUNDWATER INFILTRATION: LABORATORY COLUMN STUDIES ELMAR KUHN, MARK VAN LOOSDRECHT*, WALTER GIGER and I~N~ P. SCHWARZENBACHt Swiss Federal Institute for Water Resources and Water Pollution Control (EAWAG), CH-8600 Diibendorf, Switzerland

(Received September 1986) Abstract--In laboratory column experiments with aquifer material collected from a natural fiver

water/groundwater infiltration site, the effects of changes in NTA concentration (0.06-3.40/~M), temperature (5-20°C), and redox conditions on the microbial degradation of NTA during infiltration have been investigated. Under both aerobic and denitrifying conditions, NTA was rapidly mineralized and supported microbial growth as a sole carbon and energy source. The presence of other degradable organic compounds and of trace metals had no significant effect on the total rate of NTA elimination after a 21.8 cm flow distance. At concentrations between 0.02 and 0.05/~M, NTA degradation was still rapid (apparent pseudo first-order rate constants of up to 15 d- ~). From the results of the column experiments it may be concluded that under environmental conditions typical for Switzerland, very low residual NTA concentrations (< 0.01 ~uM) should be present at all times of the year in the groundwater after only a few meters of infiltration, even when concentrations of NTA in river water reach 3-4/~ M. This conclusion is corroborated by results of field measurements. Key words--nitrilotriacetate (NTA), groundwater, infiltration, biodegradation

INTRODUCTION

Due to the increasing use of nitrilotriacetate (NTA) as a phosphate substitute (builder) in detergents, the environmental exposure and risk assessment of this complexing agent has drawn considerable attention. This is reflected in a vast literature dealing with the behavior and effects of NTA in aquatic systems (Bernhardt, 1984; Perry et al., 1984; Anderson et al., 1985). There are, however, still some important gaps in our knowledge concerning the environmental fate of NTA, in particular with respect to its microbial degradation in natural waters. Only very few studies have investigated the behavior of NTA during river water/groundwater infiltration (Hrubec and van Delft, 1981). This aspect is of great interest, since river water infiltration is an important process in the natural recharge of groundwater, and an increasing number of waterworks use bank or dune filtration as a first step in water treatment. In a natural river water/groundwater infiltration system, important environmental conditions such as temperature, pH, redox conditions, and substrate concentrations may vary considerably with both time and space. Although there is strong evidence that, at least under aerobic conditions, NTA is readily de-

*Present address: Department of Microbiology, Agricultural University, NL-6703 CT Wageningen, The Netherlands. tAuthor to whom all correspondence should be addressed.

graded in the subsurface environment (Dunlap et al., 1972; Tiedje and Mason, 1974; Hrubec and van Delft, 1981; Larson and Ventullo, 1983; Back and Clesceri, 1986; Ward, 1986), not much quantitative data on the influence of environmental factors on NTA degradation during infiltration and in groundwater are available. For example, there exists only very limited information on the rates of anaerobic degradation of NTA under natural conditions (Tabatabai and Bremner, 1975; Ward, 1986) and on the degradation of NTA at low temperatures and at very low concentrations (<0.3/zM). The existence of a possible "threshold" concentration (Alexander, 1985), below which no further NTA elimination may occur during infiltration, is a major concern since, in some countries, the maximum NTA concentrations suggested for drinking water are as low as 0.015/~M (Kanowski, 1986). In a recent year-round field investigation at a natural river water/groundwater infiltration site in the lower Glatt Valley, Switzerland, Schaffner et al. (1986) found that NTA was always eliminated to concentrations <0.01/~M during the first 7 m of infiltration even at temperatures below 5°C (Fig. 1). During the summer months at this site, denitrification and manganese reduction are observed between the river and the first observation well (Schwarzenbach et al., 1983). The elimination of NTA during infiltration was postulated to be mediated by biological processes, and laboratory column studies were initiated to test this hypothesis and to investigate the microbial metabolism of NTA under various conditions typical

1237

1238

ELMAR KurtN et al.

for the infiltration site. T o this end, l a b o r a t o r y colu m n s were filled with aquifer material t a k e n from the field site in the G l a t t Valley a n d continuously operated u n d e r saturated-flow conditions. The columns were fitted with a series o f sampling ports, thus allowing analyses to be m a d e over the length of the column. In a previous study, this experimental setup was found to be a very useful tool for investigating the t r a n s p o r t a n d t r a n s f o r m a t i o n b e h a v i o r o f trace organic c o m p o u n d s u n d e r simulated saturated g r o u n d w a t e r flow conditions ( K u h n et al., 1985). In this paper, we report the result of experiments in which the microbial m e t a b o l i s m of N T A in the columns was studied u n d e r b o t h aerobic and denitrifying conditions. The effects o f changes in N T A concentration, temperature, a n d o f the presence of other substrates on N T A d e g r a d a t i o n were investigated.

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EXPERIMENTAL Chemicals and solutions

NTA (acid form) was of highest available purity (Fluka, Buchs, Switzerland) and was used as received. Uniformly labeled [14C]NTA with a radiochemical purity of 98% was obtained from Procter and Gamble (Cincinnati, Ohio). In order to simulate the conditions at the field site as closely as possible, a synthetic river water for aerobic experiments and a synthetic groundwater for anaerobic experiments were prepared as follows. Aqueous solutions (double distilled water) containing 0.77 mM Na +, 0.13 mM K +, 0.62 mM Mg 2+, 0.35raM NO~-, 0.03 mM HPO~-, 0.21 mM SO~4-, and 1.34 mM CI- were placed in a bottle with an excess of granulated marble (Merck, Darmstadt, F.R.G.) and while stirring, were bubbled with either air for aerobic experiments or with nitrogen gas containing 0.5% CO2 for anaerobic experiments. Equilibrium was obtained after about 24 h as indicated by a constant pH of 8.3 and 7.4, respectively. These solutions are referred to herein as carbonate media. For the experiments with actual river water, water from the River Glatt was filtered through a 0.2 g m membrane filter (Acrodisk, Scan AG, Basel, Switzerland), autoclaved, and equilibrated over CaCO3 as described previously. The major ion composition of the river water was similar to that of the synthetic river water (see above), except for phosphate which was only 0.001 mM, and it contained 6 m g l - ' of dissolved organic carbon. The low phosphate concentration was probably due to precipitation of calcium phosphate during autoclaving. Column design and column packings

The columns (25 x 5 cm i.d.) were constructed of transparent Plexiglass (acrylic tubing). The sampling ports were Plexiglass stoppers fitted with stainless steel capillary tubes with headpieces of sintered stainless steel positioned in the center of the columns. The first sampling port was located at a distance of 1.8 cm from the column inlet with four additional ports at intervals of 5 cm. The columns were filled with aquifer material from the interface of the fiver water/groundwater infiltration site in the lower Glatt Valley (Hoehn et al., 1983). The material which consisted mostly of old river sediment was air-dried, sieved to pass a mesh of 63-500/~m, and then packed into the columns. The organic carbon content of the dried sediment was 3.6% as determined by the method of Baccini et al. (1982). From analysis of the breakthrough curves (Bear, 1979) using sodium chloride as a conservative tracer,

,

,T

8 distance time

. . . . 10 [m]

,;

12

Idays}

Fig. 1. Concentration ranges ( - - - - ) and average concentrations ( ) of NTA in the River Glatt and in the

infiltrated river water at various distances from the river (data from Schaffner et al., 1986).

porosities of approx. 0.40 and hydrodynamic dispersivities on the order of 0.1 era were calculated for all columns. Experimental setup Figure 2 shows the experimental setup used for all experiments. Carbonate medium (columns I and II) or water from the River Glatt (column III) supplemented with NTA was placed in the reservoir and continuously bubbled with air or with the NJCO2 mixture described above. The medium was pumped through the columns by a high-performance liquid chromatography (HPLC) pump (Kontron, Zurich, Switzerland). The columns were operated in the saturated, upflow mode. The average velocity of the liquid flowing through the column was 8.0em h - ' (flow rate = 1 ml min-l), a velocity which is typical for the flow velocities found in the Glatt River infiltration zone (Hoehn et al., 1983). Hence, the average residence time of the water in the column was approx. 3 h. In order to prevent oxygen intrusion during experiments run under anaerobic conditions, all parts of the system were constructed of either glass, Plexiglass, or stainless steel. Bacterial filters (0.2 #M) prevented contamination of the reservoir. All experiments were conducted as step-input experiments; that is, the input concentrations of NTA, which ranged from 0.06 to 3.4#M, were kept constant for a specific period of time. The columns were operated in thermostatically-controlled rooms in the dark. If not specifically indicated, the experiments were conducted at 20°C which is the water temperature at the field site during the summer months during which anaerobic conditions were observed in the near field of the river. Analytical methods The samples from the influent and from the different sampling ports were taken at controlled flow rates of 0.1 mlmin -~ (10% of the flow rate through the column) using a peristaltic pump. For preservation, an aliquot of a 37% (w/w) formaldehyde solution was added to the sampling tube (to yield a final concentration of about 1%). The samples (typically 5 ml, and 100 ml for measurements of very low concentrations) were stored in polyethylene vials at 4°C in the dark. The NTA concentrations were determined by the method of Schaffner and Giger (1984). After

Microbial degradation of NTA Activated

carbonfilter +bacterialfilter

1239

Bacterial filter .

I

T

T

HPLC

t/)

pump

es E t~ to ¢q

cp =1

E

, xm

/

E

CaC03

(..O

Influent samplingport

Reservoir

Fig. 2. Experimental setup of the saturated-upflow sediment columns.

a concentration step on an anion exchange resin (Dowex 1X-2, 50-100 mesh), NTA was esterified with butanol/HCl. The resulting tributyl ester was then determined by glass capillary gas chromatography using a nitrogen specific detector. The detection limits were 0.02 # M for 5 ml samples and 0.001 p M for 100ml samples. For determinations of ammonium, nitrite, nitrate, and dissolved organic carbon (DOC), an aliquot of the sample was passed through a 0.2 p m membrane filter immediately after sampling, and the sample was stored at 4°C in the dark. The concentrations of these chemical species were determined by Standard Methods [APHA, 1980 (NH +, NOr, N O ; )]. DOC was determined by peroxydisulfate/u.v. oxidation and subsequent CO2 measurement by a carbon analyzer (Dohrmann DC 801). The detection limit was 0.1 mg C 1-I. Dissolved oxygen was routinely monitored in a flowthrough microcuvette with an oxygen electrode (Type IL 213, W. Ingold AG, Zurich, Switzerland). During the aerobic experiments, the influent solution was equilibrated with air (see above) resulting in a mean oxygen concentration of 0.29 mM O: at 20°C. In the experiments with the carbonate media, the effluent concentrations were always >0.10mM 02. The detection limit was 0.001 mM 02.

Experiments with 14C-labeledNTA Experiments with uniformly labeled [~4C]NTA were conducted under both aerobic and anaerobic conditions. During these experiments, the [t4C]NTA was added to the influent for a certain period of time (typically during 12-24h). The activity in the influent was approx. 20,000 dpm ml-L The collection and fractionation of the '4C-labeled chemical species in the samples from the various sampling ports was performed as follows. The sample (typically 10 ml) was passed through a 0.2 p m membrane filter to collect any ~4C-activity associated with suspended particles (i.e. suspended biomass). The filtrate was placed in a glass tube containing I ml of IM H2SO4. This acidified sample was then purged with nitrogen (30 ml min -~) for 1 h. The gas stream was passed through two tubes, each containing 10 ml of 2M NaOH to trap the 14CO2, and subsequently through a tube with 10 ml hexane/isopropranol (v/v: I/I) to trap any volatile '4C-containing organic compounds ("volatile" fraction). The ~4C-activity remaining in the acid solution is referred to as the "acidic" fraction. After washing the membrane filter twice with 10ml carbonate medium to remove any soluble '4C, the filter was incubated with I ml Solutron (Kontron, Zurich, Switzerland) for 5 h at 50°C. Ten milliliters of Lipotron (Kontron,

Zurich, Switzerland) were then added, and the ~4C-activity was determined with a scintillation counter (Packard, Tricarb Liquid Scintillation Spectrometer 3255). The activities in the various glass tubes (acidified and purged sample, NaOH-solutions, hexane/isopropanol mixture) were determined by mixing 1 ml liquid from a given tube with 10 ml Kontrogel and subsequent analysis with the scintillation counter. For analysis of the ~4C-label incorporated it/t0 th.d biomass in the columns, the columns were opened and samples of aquifer material were taken at various distances from the inlet. These samples (typically 1 g wet wt) were washed twice with l0 ml carbonate medium by centrifuging, and acidified with IM H2SO4 to remove any inorganic 14C. The samples (typically 50 mg dry solid) were then processed as already described for the samples collected on the membrane filters. RESULTS

Experiments under aerobic conditions Breakthrough and initial degradation. Figure 3 shows the b r e a k t h r o u g h b e h a v i o r o f N a C I a n d N T A after a flow distance o f 21.8 cm in two columns which were o p e r a t e d with c a r b o n a t e m e d i a u n d e r identical conditions. T h e o r d i n a t e in Fig. 3 (as in m o s t o f the following figures) is the ratio o f the c o n c e n t r a t i o n at

............ 1.0-

= C0[umn [

oColumn~.

/

~

f'

NaCI ~

~

NTA

0.75" 0.50.25" I

o

o

;

f~"

2

Time

~

~

[hours]

Fig. 3. Breakthrough behavior of NaC1 ( - - - - ) and NTA in column I ( 0 ) and column II (O) after 21.8 crn flow distance. The breakthrough curve for NaC1 was the same in both columns. The influent concentration was 3.4 mM for NaCI and 0.55 ltM for NTA.

ELMAR KUrtN et al.

1240

a given sampling port (C) to the input concentration (Co). From the breakthrough curves in Fig. 3, it can be seen that NTA was only very slightly retained in the columns; a retardation factor of < 1.2 is obtained when assuming conservative behavior of the NaC1. The fact that NTA was in effect not sorbed by the aquifer material used is not too surprising since, under the conditions of this experiment (no trace metals added), it may be assumed that NTA was present in solution predominantly as negatively charged hydrophilic calcium and to a lesser extent as a magnesium complex, (i.e. [CaNTA] , [MgNTA] ). Furthermore, Hassett and Means (1980) have demonstrated that the presence of organic material in natural sorbents strongly reduces the sorption of hydrophilic compounds capable of forming surface complexes with mineral surfaces. This might explain the significantly weaker sorption of NTA by our organic carbon-rich aquifer material as compared with various soils exhibiting much lower organic carbon contents (Baek and Clesceri, 1986). The breakthrough curve of NTA in Fig. 3 shows that, in contrast to the NaCI, the effluent NTA concentration never reached the input concentrations but leveled off after 4 h at about 0.9C0 in both columns. Thus, some NTA elimination had already taken place during the first hours of operation. For a quantitative measure of the elimination of NTA in the columns, we can define an apparent pseudo first-order rate constant, k~, for a given column segment: In (C2/CI) ko - - AL/u

(1)

where Ct and C2 are the NTA concentrations at the beginning and end of the column segment, AL is the flow distance, and u is the average flow velocity. On the first day of the experiment, one obtains a k~ value for the entire flow distance (0-21.8cm) of approx. 1 d-L for both columns.

Microbial adaptation and effect of concentration on NTA degradation. Figure 4a shows the time course of the relative NTA concentrations (C/Co) at two sampling ports (1.8 and 21.8 cm) during the first month of operation. Figures 4b and c show concentration profiles over the whole column length at selected days. The data shown are from column I; any differences observed between the two columns are addressed below. During the first 13 days, the influent NTA concentration was 0.55 pM. As can be seen from Figs 4a and b, the rate of NTA elimination increased substantially over this time period. The k~ values calculated for both columns at various days for the entire column length are summarized graphically in Fig. 5. When considering NTA elimination over the whole column, it can be seen that until the 6th day the k~ values doubled at a more or less constant rate (doubling time, t2 = 1.5 and 1.7 d, respectively). After day 6, after which most of the NTA consumption

already occurred in the first third of the column, the increase in ka for the whole column slowed down significantly (t 2 = 5 and 6 d, respectively). Over the first segment, however, ko continued to increase with a doubling time of approx. 2 days. While columns I and II gave very similiar results when considering the elimination of NTA over the whole column length, some small differences were observed between the first segments (0.1 1.8 cm). The initial activity in the first 1.8 cm of column I was somewhat higher than in column II; the t, value, on the other hand, was somewhat lower. On day 13, the effluent concentrations in both columns at 21.8 cm were <0.02 pM; that is, over 95% of the NTA was eliminated. The k,~ values were approx. 6 0 d ~ for the first segment ((~ 1.8 cm) and 30 d ~ for the whole column. At this time the influent concentration was reduced by a factor of 10 to 0.06/aM. The response of column I upon reduction of the influent concentration is shown in Fig. 4a (column II exhibited a very similar behavior). While the ka value for the first segment dropped by approximately a factor of 6, it decreased in the latter part of the column (!.8-21.8 cm) by only a factor of 2. After 4 days of operation with an influent concentration of 0.06/aM (during which the elimination rates of NTA increased only very slightly in all parts of the columns), the k,, values for the entire column (as well as for the first segment) were approx. 15 d and the effluent concentration at 21.8cm was 0.01 pM. C,~ was then again increased to 0.84#M (day 18). Only one day alter changing the influent concentration to 0.84stM, the k~-values for the entire column had resumed the value previously found (day 13) at a similar concentration. On day 22, the influent concentration was then raised to 3.14pM. This increase in concentration initially led to a decrease in k,~ by a factor of two for the whole column, which was mostly due to the almost 10-fold decrease in k~ in the first segment. On day 23, the NTA concentration in the effluent at 21.8 cm was 0.40 pM. During the next 5 days, however, similar to what was observed during the initial phase of the experiment, the k~ value for the first segment doubled at a constant rate with a doubling time of about 1.5 days. On day 31, >95% of the NTA was eliminated over the first 1.8 cm (see profile in Fig. 4c). The k~ for this segment had increased to about 340 d- ~, the ko for the remainder of the column (1.8-21.8 cm) was approx. 15 d -~, and the NTA concentration in the effluent at 21.8 cm was 0.03 pM. It should be pointed out that after 1 month of operation, both columns gave approximately the same results. Column I was then used to study the influence of temperature on NTA degradation, and to establish a carbon and nitrogen balance under aerobic conditions. Column II was operated under anaerobic conditions. Effect of temperature. The effects of sudden temperature changes on NTA degradation in column I

1.8

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Fig. 4. Aerobic degradation of N T A in column I: initial phase of adaption and effect of N T A concentration changes, a--Time course of the relative effluent concentration (C/Co) of NTA after 1.8 and 21.8 cm flow distance, b, c--Concentration profiles of N T A on various days.

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1242

ELMARKUnN et al.

,.o]~ o~y 31 (2o°c)--II

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Fig. 5. Apparent pseudo first-order rate constants, ka, calculated for the elimination of NTA over the entire flow distance (0-21.8 cm) in columns I (0) and II (O) at various days during the first 2 weeks of operation under aerobic conditions. The infiuent NTA concentration was 0.55 izM.

~;.8

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distance 0ay

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o.25t ~\ are summarized in Fig. 6. As can be seen from Fig. 6a, a decrease in temperature from 20 to 10°C led to a temporary decrease in activity in the first segment of the column where the k a value dropped by a factor of two from 340 to 170d -1. In contrast, for the remainder of the column (1.8 - 21.8 cm), ka increased by a factor of about 1.5. Note that due to the decrease in activity in the first segment, the organisms in the subsequent part of the column were exposed to higher NTA concentrations. During the next 5 days, the activity in the first segment steadily increased and on day 36, virtually the same profile was found at 10°C (Fig. 6b) as was detected on day 31 at 20°C (Fig. 6a). The column was then transferred to a 5°C thermostatically-controlled room. The effects found when lowering the temperature from 10 to 5°C were qualitatively the same as observed for the temperature change from 20 to 10°C; that is, a significant decrease in k~ for the first segment, but a slight increase in k a for the remainder of the column (Fig. 6b). However, it should be noted that lowering the temperature by 5°C from 10 to 5°C caused a larger decline in ka for the first segment (factor 2.5) than lowering the temperature by 10°C from 20 to 10°C (factor 2). During the next 9 days at 5°C, the effluent concentration at 21.8cm stayed constant at 0.15 I~M, although the k a value for the first segment of the column increased during this time by nearly a factor of 2 (Fig. 6c, day 46). This increase was balanced by a simultaneous decrease in k~ in the latter part of the column. On day 46, the column was brought back to the 20°C room. Immediately, the ko value for the first segment increased by a factor of 2, and a very similar profile as measured on day 31 was obtained (cf. Figs 6c and a).

6.8

o

~

o[~°! ~ 1.8

,

t

6.8

11".8 Flow distance

16.8

It

21.8

{cm]

Fig. 6. Concentration profiles of NTA in column I (aerobic conditions) immediately before and after temperature was changed from a--20 to 10°C; b ~ t 0 to 5°C; c--5 to 20°C. The influent concentration was 3.14 #M.

Carbon and nitrogen balance. Between day 47 and 52, no nitrogen sources other than NTA (3.14 laM N) were present in the influent, which had no detectable influence on the NTA concentration profiles. However, after a flow distance of 21.8 cm nitrate was found at a concentration of 4.3 + 1.2 #M. Since no ammonium or nitrite were detected, it is most likely that the nitrogen liberated from NTA as ammonium was oxidized to nitrate by nitrifying bacteria. On day 52, ~4C-labeled NTA was added to the influent of the column for 12 h. Figure 7 shows the profiles of ~4CO2 and J4C-activity measured in the acidic fraction (see "Experimental" section), expressed as the percentage of ~4C-NTA activity in the influent. The profiles of ~4C-activity in the acidic fraction were identical with the NTA profile suggesting that all activity in this fraction was undegraded labeled NTA. No activity was detected in the volatile fraction, and < 0.1% of the total activity was recovered from the membrane filters (particulate fraction). From the total activity introduced, about 68% was recovered as ~4CO2 at the first sampling port. The slight decrease in 14CO2-activity found over the length of the column was probably due to incorporation of labeled carbonate into the solid matrix which consisted of approx. 30% calcite (Hoehn et al.,

1243

Microbial degradation of NTA

,J,L ....p__,i.8 14C-activity

i

in attached

biomass

2:.8

6.8

Ia

Flow

11.8 distance

Icml

Fig. 7. Results of experiments with “C-labeled NTA under aerobic conditions. The influent NTA concentration was 3.14 pM. The “‘CO2 and “C recovered in the acidic fraction are expressed as percent of the “C-activity in the influent. The dashed lines indicate the 14Crecovered in attached biomass.

1983) and/or due to 14C02 fixation by autotrophic bacteria. After addition of 14C-labeled NTA over 12 h, the column was opened and the activity of the attached biomass was determined at various distances from the inlet. A composite sample was taken for the first segment of the column, and discrete samples were taken at the sampling sampling ports. Figure 7 shows that the highest biomass 14C-activity was found in the first segment where most of the NTA elimination occurred. Integration over the whole column length

indicates that approx. 15% of the total L4C-activity was incorporated into biomass. Experiments under anaerobic conditions

Figure 8a shows the time course of the relative NTA concentrations at the 1.8 and 21.8 cm sampling ports upon removal of oxygen from the influent of column II. The influent concentration of NTA was 3.40 PM. Note that upon removal of oxygen the time is reset to zero. The removal of oxygen initially led to a significant decrease in NTA elimination, particu-

Time

W3mM

[days

1

I

c Aerobic

’

-I

Denllrlfying

02)

added

(0.35

k

mM NOj)

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Removal

Day

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Fig. 8. Anaerobic degradation of NTA in column II. The influent NTA concentration was 3.4OpM. a-Time course of the relative effluent concentration of NTA (sampling ports 1.8 and 21.8 cm) after removal of molecular oxygen from the influent. Response to the removal and addition of nitrate in the influent. b-Concentration profiles of NTA at various days during the first 2 weeks upon onset of anaerobic conditions. c-Concentration profiles of NTA, nitrate, and nitrite on day 45.

1244

ELMARKUHNet al.

larly in the first segment of the column. On the 1st day under anaerobic conditions, the k~ value for the whole column was 3d-~; that is, about 10 times smaller than on the previous day under aerobic conditions. However, within 2 days, the k~ for the whole column had resumed a value of about 20 d and then increased over the next 2 weeks to 40 d-J (see profiles in Fig. 8b). In the first segment (0-l.8 cm), k~ increased between day 3 and 17 from 2 to 150 d ~. It should be noted that between day 6. and 10, the concentration profile did not change (Fig. 8a). On day 17, the effluent concentration at 21.8 cm was 0.03 #M. Column II was then transferred to a 5:C thermostatically-controlled room. The effect of the temperature change from 20 to 5 C on NTA elimination in the anaerobic column is shown in Fig. 9a. In the first two segments of the column (0-1.8 and 1.8-6.8 cm), the k~ value decreased by about a factor of 4, while in the remainder of the column (6.8 21.8 cm) k~ increased. After 3 weeks at 5°C (day 37), the effluent concentration at 21.8 cm had dropped to 0.04/~M; that is, the overall elimination rate was similar to that observed 3 weeks earlier at 20°C (day 17), although the k, values for the various segments were quite different (cf. profiles in Figs 9a and b). When the column was brought back to the 20°C room, the k~ for the first segment increased by about a factor of 4, while the overall elimination rate did not change significantly. On day 45, the NTA profile in the anaerobic column II (Fig. 8c) was very similar to the profiles obtained for column I after one month of operation under aerobic conditions (see for example, NTA profile on day 31, Fig. 4c). Figure 8c shows that in the first segment, where most of the NTA elimination took place, nitrate was consumed and nitrite was produced. When nitrate was removed from the influent of the column, the NTA degradation ceased immediately, but resumed upon renewed addition of nitrate (Fig. 8a). These findings clearly demonstrate that the degradation of NTA was coupled to the reduction of nitrate. On day 50, ~4C-labeled NTA was added to the influent of column II over a period of 12 h. Very similar results as found for column I under aerobic conditions were obtained with regard to the recovery of the ~4CO2 (72%) and of t4C-label in the acidic fraction. The profile of ~4C-activity in the acidic fraction was again identical with the NTA concentration profile. No activity was found in the volatile fraction, but about 0.7 and 0.3% of the total activity were recovered in the particulate fraction at 1.5 and 6.5 cm, respectively. In contrast to the findings under aerobic conditions, only 3% of the total ~4C-labelhad been incorporated into biomass, and this again mostly in the first segment of the column. Experiments with river water

In order to evaluate the influence of other substrates on the degradation of NTA during infil-

1.0 0.75 0.5

Day 17 (20°C) ---* Day 175 (5°C)

1

0.25 ",o

0

o

1,8

6.8

11'.8 Flow distance

[cm]

~°~ Day 37 <5°c)--,. Day 37.s (2°°)

21.8

~(,.b)

0.75 ~

0.5

0.25

0

20°C ~ " i ' ~.8

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i~.8

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Fig. 9. Concentration profiles of NTA in column II (anaerobic conditions) immediately before and after the temperature was changed from a--20 to 5°C; ~ - 5 to 20°C.

tration, a third column (column III) was operated at 20°C with membrane filtered (0.2 #m) aerated water from the River Glatt. As in the initial experiments with columns I and I1, the influent NTA concentration was 0.55#M. In contrast to the previous experiments, however, the influent water contained 6 m g l -~ dissolved organic carbon (DOC) and 0.07 mM ammonium. Figure 10a shows that on the first day, elimination of NTA over the entire flow distance (0-21.8 cm) was very similar to that found for columns I and I I (C ~ 0.9 Co); however, the rate of NTA elimination in column III increased only about half as fast as in the other columns, in both the first segment as well as the whole column (t2 = 3 days). From the concentration profiles shown in Fig. 10b for day 13, it can be seen that in the first two segments (0-6.8 cm), the major portion of DOC elimination took place and that oxygen was completely consumed. In this part of the column, ammonium and nitrite were also eliminated, and nitrate was formed (Fig. 10c). On the other hand, between 6.8 and 21.8 cm, nitrate was depleted and nitrite was formed again. Thus, in column III, NTA was degraded aerobically in the first third of the column, and anaerobically under denitrifying conditions over the remaining flow distance. After 25 days of operation, the effluent concentration of NTA at 21.8 cm in column III was 0.04#M; that is, more than 90% of the NTA was eliminated over this distance. The final k~ for the whole column was about 25 d l, which is of the same magnitude as the k, value obtained after 13 days at the same NTA influent concentration for columns I and II operated with the aerobic carbonate medium containing NTA as the sole carbon source. The experiment with

Microbial degradation of NTA

10

5

15 Time

D,, 13

1245

20

~

0.)' 13

0.75-

25

[days]

(~

0.4

DOC

~-'5"

___

~

c.)

~ 0.3 0"2 I NHz~

0.25-

0.1 ~

0 1.8

6.8

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Flow

distance

16.8

21.8

[cm]

NO~

0 'r"-~-"

-7,

1.8

6.8

Ip Flow

11:8 16:8 distance [cm]

21.8

Fig. 10. Degradation of NTA using river water as influent medium (column III). a--Time course of the relative concentration of NTA after 1.8 and 21.8 cm flow distance. ~ o n c e n t r a t i o n profiles of NTA, dissolved organic carbon (DOC), and molecular oxygen on day 13. c--Concentration profiles of ammonium, nitrate, and nitrite on day 13.

t4C-labeled N T A in column III gave very similar results as with the other columns, with the exception that ~4CO2 evolution was somewhat lower and an even higher portion of the ~'C-label was incorporated into the biomass than in column I in the area of maximum N T A degradation. DISCUSSION

The observed temporal and spatial changes in the N T A degradation rates and the results from the experiments with 14C-labeled N T A clearly indicate that under both aerobic and anaerobic conditions, N T A was rapidly metabolized by microorganisms in the columns, and that N T A supported growth serving as sole carbon and energy source. Under anaerobic conditions, nitrate served as the terminal electron acceptor. Our findings are in strong support of the scarce evidence available that N T A is readily mineralized under denitrifying conditions in the subsurface environment (Tabatabai and Bremner, 1975; Ward, 1986). Under aerobic conditions, the initial rate of NTA degradation was very similar in all three columns with an apparent pseduo first-order rate constant, ka, of about 1 d -~ at 20°C. Although it is somewhat tenuous to compare this rate constant to rates of NTA elimination determined in samples from the natural aquatic environment, it is interesting to note that the reported values are all in the same order of magnitude. At concentrations comparable to our experiments, Larson et al. (1981) and Larson and Davidson (1982) reported pseudo-first order rate constants for various river waters of between 0.25 W.R. 21/lOnG

and 1.3 d -1 at temperatures between 14 and 24°C. Larson and Ventullo (1983) found rate constants of 0.31 and 0.54d -I for water samples derived from groundwater wells. The rate constants for N T A were only a factor of 4 smaller than the ones determined for a naturally occurring amino acid (glutamic acid). These findings and the results of other studies with soil samples (e.g. Tiedje and Mason, 1974; Baek and Clesceri, 1986; Ward, 1986) suggest that under aerobic conditions, N T A can be mineralized by many different microorganisms present in the natural environment. During the first 6 days of operation, the rate of N T A degradation in columns I and II (k, value) increased with a doubling time of about 1.5 days (Fig. 5). In batch cultures, Enfors and Molin (1973a) reported a generation time of 3 days for a pure culture (strain NTA-A2) growing on N T A under aerobic conditions at 10°C. Egli (EAWAG, personal communication) found doubling times on the order of 0.25-0.5 days for several different aerobic, NTAdegrading pure cultures at 30°C. From these data, we conclude that in our columns operated aerobically at 20°C, the NTA-degrading microorganisms grew exponentially in the whole column during the first days, and in the first segments for at least another week. In the latter part of the columns (6.8-21.8 cm), due to the steadily decreasing N T A concentration, growth slowed down significantly after day 6 (Fig. 4). The immediate drop of the apparent rate constant k, upon decreasing the influent N T A concentration to 0.06 # M (Fig. 4) suggests that some portion of the microorganisms in the column, in particular those in the first segments which had been previously exposed

1246

ELMAR KUH~ et al.

to higher concentrations, were unable to utilize NTA at such low concentrations. This is consistent with the findings made in the latter part of the columns (6.8-21.8 cm) where at higher influent concentrations, a drop in the ka value was observed when, due to the increasing activity in the first part of the column the NTA concentration decreased to < ~ 0.1 # M. However, it should be pointed out that even at very low concentrations, k a values of up to 15 d-~ were found, resulting in effluent NTA concentrations of as low as 0.01/~M after 21.8 cm flow distance. The observed decrease in k a (particularly in the first segment of the column) upon increasing the influent NTA concentration to a level well above the concentration to which the microorganisms were first exposed during the adaptation period (i.e. from 0.84 to 3.14/~M, see Fig. 4), was probably due to substrate saturation of the enzymes responsible for NTA degradation. A half-saturation constant as small as 0.4/~M has been reported for a NTA-degrading bacterium (Wong et al., 1973). The subsequent increase in k~ with a doubling time of 1.5 days again indicates exponential growth of the microorganisms in the column. The response of column II upon removal of oxygen (Fig. 8) suggests that at least some of the organisms present were able to degrade N T A after a very short induction period ( < 1 day) using nitrate as the electron acceptor. Enfors and Molin (1973b) have described an organism cultivated under aerobic conditions that also degraded NTA under denitrifying conditions after an adaptation phase of a few days upon removal of oxygen. Between day 6 and 10, a stagnation in activity was observed (Fig. 8a). We have observed a very similar behavior as shown for NTA in Fig. 8a in an anaerobic xylene-degrading column where, before the stagnation phase, nitrate was stoichiometrically converted to nitrite (Kuhn et al., 1985). After this phase, nitrate was also reduced to gaseous products. In column III, the presence of high concentrations of other organic substrates in the influent of the column (6 mg 1- ~dissolved organic carbon present in the river water used) led to a complete depletion of oxygen within the first 6.8 cm. In this part of the column, in particular in the first segment (0-1.8 cm), the major portion of the easily degradable DOC was eliminated (Fig. 10b). The doubling time of the rate of N T A elimination in this aerobic part of column III (h = 3 days) was twice as long as found in columns I and II where no other substrates were present in the influent. Probably, microorganisms that would also have been able to degrade NTA grew on other organic compounds present in the river water. In the anaerobic part of the column, the slower increase in activity during the first 15 days (h ~ 3 days) is consistent with the findings of Egli and Weilenmann (1986) who observed a generation time of 1 day for an NTA-degrading mixed culture under denitrifying conditions, which was about a factor of 2 larger than

the doubling times of their aerobically NTAdegrading pure cultures. Nevertheless, despite these longer adaptation times in column III, after 25 days of operation, overall NTA elimination was almost 95%. Changes in the NTA concentration profiles observed upon temperature changes under aerobic and anaerobic conditions (Figs 6 and 9) in our columns, give additional insight into the dynamics of pollutant degradation in groundwater infiltration systems where, due to continuous advective flow through the porous medium, the susbstrate availability to the (attached) microorganisms that are more remote from the source of pollution depends very much on the activity of the organisms located much closer to this source. As already mentioned, under aerobic conditions, we observed significant decreases in activity in those parts of the column where the NTA concentration dropped below 0.1 # M. Under anaerobic conditions this level was somewhat higher (0.25pM). Each time before the temperature was decreased, the NTA concentrations beyond a 6.8 cm flow distance were below this critical level in both columns. Upon decreasing the temperature, the activity of the microorganisms in the first segments decreased, which led to an increase in NTA concentration above the critical level in the latter column segments, with the consequence that here, the overall elimination rate increased instead of decreasing. Larson et al. (1981) have investigated the effect of temperature on the NTA degradation rate in river water samples. For the temperature range between 2 and 24°C, they found a temperature coefficient, Qt0 = 2; that is, a doubling of the degradation rate with a 10°C increase in temperature. This is consistent with our findings in the first segment of the column (where the NTA concentration was always above the critical concentration) when changing the temperature from 20 to 10°C. However, when lowering the temperature to 5°C, we observed a significantly greater effect. Tiedje and Mason (1974) demonstrated for NTA degradation in soils, that when lowering the temperature to 2°C, the elimination rate of NTA depended very much on the previous acclimation temperature. They found a much higher activity at 2°C when the acclimatization temperature was 12.5°C as compared with 24°C. In a natural river water/groundwater infiltration system, temperature changes occur slowly (not abruptly as in our experiments), and the microorganisms may adapt even better to the low temperatures. Thus, as shown by the NTA profiles obtained after a certain acclimation period at 5°C, under both aerobic (Fig. 6c) and denitrifying (Fig. 9b) conditions, it can be expected that NTA is very efficiently eliminated during infiltration even at very low temperatures. Finally, the effect of metal ions on the speciation and thus on the degradation behavior of NTA during infiltration should be shortly addressed. Madsen and Alexander (1985) have recently shown that depending on the metal ions present, various complexing agents

Microbial degradation of NTA including NTA were mineralized aerobically at different rates by both sewage microorganisms and pure cultures. They concluded that, in the case of NTA, only the calcium complex was readily degraded. On the other hand, Tiedje and Mason (1974) found rapid mineralization of various NTA metal complexes added to soils. They argued that in the presence of the soils used (pH 5-7), due to adsorption of the trace metals, NTA would be predominantly present as iron and calcium complexes, both of which they found to be readily degradable. In any case, when comparing our results obtained from the experiments with carbonate media where no trace metals were added to the influent (columns I and II) with those from column III which was operated with water from the quite heavily polluted River Glatt (for typical metal concentrations see von Gunten and KuU, 1986), no obvious effects of the heavy metals present in the river water on NTA degradation were detected. Environmental significance

The results of our laboratory column studies indicate that during river water/groundwater infiltration, NTA is readily mineralized by microorganisms under both aerobic and denitrifying conditions. Due to the spatial distribution of the microorganisms along the infiltration path and due to their ability to adapt rapidly to changes in temperature, NTA concentration and redox conditions (aerobic, denitrifying), it can be expected that under the conditions typical for most infiltration sites in Switzerland, very low residual NTA concentrations (<0.01/~M) will be found in the groundwater even after only a few meters of infiltration. At two field sites at the Glatt River (Schaffner et al., 1986) and at the Sitter River (Leuenberger and Giger, personal communication), at river water NTA concentrations in the range of 0.03-0.42#M, the NTA concentration in the infiltrated groundwater was always < 0.01 # M. These findings are consistent with the observations in our columns where NTA was still quite rapidly mineralized at concentrations well below 0.05 #M (k a values of up to 15 d-1). Furthermore, even significant increases in NTA concentrations in river waters should not lead to higher concentrations in the groundwater, since, due to microbial growth, the major part of NTA degradation will take place during the first meter, maybe even during the first few centimeters of infiltration. Consequently, a significant remobilization and transport of trace metals by NTA from river sediments to groundwater is rather unlikely. Since there is also no evidence for the formation of toxic metabolites during NTA degradation, it seems safe to conclude from our study that, under environmental conditions as they prevail in Switzerland (a.o., waters rich in calcium carbonate), even at elevated NTA concentrations in rivers of up to 3--4#M, NTA should not cause any severe problems with respect to groundwater contamination

1247

from infiltrating river water. Similar conclusions might be drawn for other sources of NTA in the subsurface such as, for example, septic tanks, at least in environments rich in calcium. Acknowledgements--This work was supported by the Swiss National Science Foundation (Project 3.944-0.84), by the Swiss Federal Institute of Technology, and by the Swiss Environmental Protection Agency (BUS). We thank C. Schaffner, F. Stauffer, and P. Adank for experimental assistance, and Procter and Gamble for providing the t4C-NTA. We are indepted to P. C. Colberg-Swoboda and J. Zcyer for reviewing the manuscript. Valuable comments were made by T. Egli. We are grateful to H. Bolliger for drafting the figures and S. Betschmann for preparing the text.

REFERENCES

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infiltration of river water to groundwater: laboratory column studies. Envir. Sci. Technol 19, 961468. Larson R. J., Clinckemaillie G. G. and Van Belle L. (1981) Effect of temperature and dissolved oxygen on biodegradation of nitrilotriacetate. War. Res. 15, 615 620. Larson R. J. and Davidson D. H. (1982) Acclimation to and biodegradation of nitrilotriacetate (NTA) at trace concentrations in natural waters. Wat. Res. 16, 1597-1604.. Larson R. J. and Ventullo R. M. (1983) Biodegradation potential of groundwater bacteria. Proc. 3rd natn. Symp. on Aquifer Restoration and Groundwater Monitoring, pp. 402-409. National Well Water Association, Worthington, Ohio. Madsen E. L. and Alexander M. (1985) Effects of chemical speciation on the mineralization of organic compounds by microorganisms. Appl. envir. Microbiol. 50, 342 349. Perry R., Kirk P. W. W., Stephenson T. and Lester J. N. (1984) Environmental aspects of the use of NTA as a detergent builder. War. Res. 18, 255 276. Schaffner C.. Ahel M. and Giger W. (1986) Behaviour of organic micropollutants during infiltration of river water into groundwater. In Organic" Micropollutants in the Aquatic" Environment (Edited by Angeletti G. and Bjorseth A.), pp. 455-459. Reidel, Dordrecht, The Netherlands.

Schaffner C. and Giger W. (1984) Determination of nitrilotriacetic acid in water by high-resolution gas chromatography. J. Chromat. 312, 413-421. Schwarzenbach R. P., Giger W., Hoehn E. and Schneider J. K. (1983) Behaviour of organic compounds during infiltration of river water to groundwater. Field studies. Envir. Sci. Technol. 17, 472-479. Tabatabai M. A. and Bremner J. M. (1975) Decomposition of nitrilotriacetate (NTA) in soils. Soil Biol. Biochem. 7, 103 106. Tiedje J. M. and Mason B. B. (1974) Biodegradation of nitrilotriacetate (NTA) in soils. Soil Sci. Soc. Am. Proc. 38, 278-283. Von Gunten H. R. and Kull T. P. (1986) Infiltration of inorganic compounds from the Glatt River, Switzerland, into a groundwater aquifer. Wat. Air Soil Pollut. 29, 333-346. Ward T. E. (1986) Aerobic and anaerobic biodegradation of nitrilotriacetate in subsurface soils. Ecotox. envir. Saf. 11, 112--125. Wong P. T. S., Liu D. and McGirr D. J. (1973) Mechanism of NTA degradation by a bacterial mutant. War. Res. 7, 1367 1374.