~ Pergamon 0043-1354(93)E0039-U

War. Res. Vol. 28, No. 8, pp. 1827-1833, 1994 Elsevier Science Ltd. Printed in Great Britain

BEHAVIOR OF NITROGEN-SUBSTITUTED N A P H T H A L E N E S IN F L OODED SOIL- -PART II. EFFECT

OF BIOAVAILABILITY ON BIODEGRATION KINETICS

BILAL AL-BASHIR [, JALAL HAWARI 2., RI~JEAN SAMSON 2 a n d ROLAND LEDUC 3

I Department of Civil Engineering and Applied Mechanics, McGill University, Montreal, Quebec, Canada H3A 2K6, 2Environmental Engineering Group, Biotechnology Research Institute, National Research Council Canada, Montreal, Quebec, Canada H4P 2R2 and 3Department of Civil Engineering, University

of Sherbrooke, Sherbrooke, Quebec, Canada JIK 2RI

(First received February 1993; accepted in revised form November 1993)

Abstract--Mineralization of l-aminonaphthalene, 2-aminonaphthalene and I-amino-2-methyl-naph- thalene under aerobic conditions in flooded soil was found to proceed with a biphasic pattern, an initial fast phase followed by a slower second one. Also the sorption isotherms of these substrates were found to be hyperbolic and were best described by the Langmuir model. When initial mineralization rates were expressed in terms of initial aqueous-phase concentrations, they gave rise to simple hyperbolic kinetics that obeyed Michaelis-Menten model for enzyme-catalysed reactions. These initial mineralization rates were found to be directly proportional to the substrate aqueous concentration reaching their maxima at about 100 #g g-t (aminonaphthalene/soil slurry). Whereas the second phase mineralization rates were found to be first order with respect to the adsorbed fraction of the substrate and showed no sign of saturation, thus indicating that biodegradation is controlled by the rate of desorption.

Key words--aminonaphthalene, mineralization, sorption, bioavailability, Michaelis-Menten, first-order kinetics.

INTRODUCTION

Biodegradation kinetics of pollutants in the soil environment provides a basis for a better understanding of their fate, persistence and potential threat to living organisms. Several studies have shown that bioavailability plays a crucial role in determining the biodegradation of various organic contaminants and several models have been devel- oped to understand biodegradation kinetics. Scow et al. (1986) have reported that the mineralization of the aromatic amine, aniline, in soil is biphasic and is attributed to several factors, i.e., the presence of two states of the substrate bioavailability, the involve- ment of different populations of organisms, or the accumulation of intermediate metabolites. Conse- quently, a two-compartment model has been devel- oped (Scow et al. 1986) to describe the biphasic phenomena encountered in the mineralization of aniline.

Others (Steen et al. 1980) have modified a second- order model, applicable to biodegradation of pollutants in natural waters, by incorporating a partition coefficient in it. The model assumes that adsorption affects biodegradation solely by decreas- ing concentration of the pollutants in the aqueous

*Author to whom all correspondence should be addressed.

phase, i.e. only the aqueous fraction is bioavailable and the adsorbed fraction is totally unavailable. Also, Hamaker (1972) proposed a biodegradation model based on simple hyperbolic kinetics and on the assumption that soil contains both active and inactive biological sites. Accordingly, the rate of degradation is proportional to the amount of contaminant adsorbed on the active sites and adsorption on active and inactive sites is proportional to the concentration of the soluble substrate.

Mihelcic and Luthy (1991) have reported that mineralization of naphthalene in a soil-water suspen- sion under denitrifying conditions is dependent on solute partitioning between soil and water. They described the overall change of soluble naphthalene in the aqueous phase by developing a model which combines Michaelis-Menten kinetics with those of sorption/desorption in soil. The model assumes that the soluble and the sorbed substrate were in equilibrium throughout the course of the experiment and suggests that complete mineralization of the adsorbed fraction is attainable. The latter two as- sumptions deal with a specific case in which desorp- tion is instantaneous and shows no hysteresis, i.e. desorption is first order with respect to aqueous concentration, and does not take into consideration the limiting case when desorption turns zero order with respect to aqueous concentration.

1827

I828 BILAL AL-BASHIR el al.

These previous studies indicate that there is a need to further investigate the extent to which bioavailabil- it5, affects biodegradat ion kinetics. Fur thermore , studies on the behavior of amino-subst i tu ted PAHs are seriously lacking in the literature. In the preceding paper (Al-Bashir et al. 1994), the mineral izat ion of three aminonaphtha lenes , i.e., 1-aminonaphthalene, 2 -aminonaphtha lene and I -amino-2-methyl-naph- thalene, was found to be biphasic and it was suggested that bioavailabili ty considerat ions were responsible for such behavior. It is the purpose of this study to investigate the kinetics of mineral izat ion of these compounds and to examine the mechanisms and the extent to which bioavailabili ty affects their kinetics.

MATERIALS AND METHODS

The aminonaphthalenes, l-amino-[l-J4C]-naphthalene and 2-amino-[8-14C]-naphthalene were prepared from a mixture of their corresponding radiolabelled and unmarked naphthols (Sigma Chemical Co., St. Louis, Mo, U.S.A.) using the Bucherer reaction (Vogel, 1978). The compounds had a specific activity of 32,874 and 36,283 dpmmg -l, respectively, and a final purity of 99 _+ %. [8-14C]-l-amino-2- methytnaphthalene was prepared by reducing [8-~4C]-l-ni- tro-2-methylnaphthatene using Urushibara catalysts (Hata, 1971) which, in turn, was prepared by direct nitration of [8>4C]-2-methylnaphthalene. For more details on the synthesis of these compounds see A1-Bashir et al. (1994).

The soil sample was obtained from Jarry Park, Montreal, Quebec and was characterized as clayey loam with soil organic matter~ organic carbon, pH and cation exchange capacity as 3.96%, 1.60%, 7.27 and 19.94me/ lOOg soil, respectively.

Varying initial concentrations of each compound were investigated. For 1-aminonaphthalene, these were, 5, 10, 20, 30, 40, 50, 100 and 150 Hg l-aminonaphthalene g ~ of soil slurry. The initial concentrations for 2-aminonaphthalene were 5, 15, 30, 50, 100 and 150~gg ~ and for l-amino-2- methylnaphthalene were 10, 20, 35, 50, 100 and 200,ug g- ~. The amino-compounds were added as a methanol stock solution and the total volume of methanol added was constant at 200t~I per vial.

Batch equilibrated adsorption experiments were con- ducted for autoclaved slurry samples at pH 6.5 following the procedure described previously (AI Bashir et al., 1990). Biodegradation was carried out in 100-ml glass vials con- taining 30 g soil slurry (30% w/w soil/water) and housing a KOH trap. The aqueous phase was a mineral growth medium prepared by the method of Thomas et al. (1986). Biodegradation was carried out under pure oxygen and the pH was kept between 6.5 and 7.0 using HC1 (1 N). Pure oxygen was also replenished every sampling session. For each concentration, six replicates were prepared of which three were analysed regularly for ~4CO2 in the alkali trap without acidification, two were sacrificed by acidification to make corrections for any CO, in the soluble form and the sixth one acted as control by receiving 500 pg/g HgCI> The evolved ~4CO2 was captured in KOH traps comprising of 10-ml glass tubes that shared the same head space with the slurry. Radioactivity in the alkali was measured using a scintillation counter (Packard Tri-Carb #4530, Downers, ILL, U.S.A.). On few occasions, duplicate samples of 2-ml slurry were taken from the biologically-active vials. These were centrifuged at 15,600 × g for 10 min (Centrifuge 5414, Eppendorff, Hamburg). The supernatant was then analysed for radioactivity.

RESULTS

The effect of initial N-subst i tuted naphtha lene concent ra t ion on the mineral izat ion of l - aminonaph- thalene, 2 -aminonaphtha lene and I-amino-2-methyl- naphtha lene is shown in Fig. 1. Concent ra t ions reported in the legend of Fig. 1 represent total initial concent ra t ion of /~g of compound per g of soil slurry. At relatively low concentrat ions, no lag period was observed for any of the studied compounds. However, a lag period became more pronounced for bo th 1- and 2-aminonaphtha lene at 1 5 0 # g g ~ and for l -amino-2-methyl -naphtha lene at 2 0 0 ~ g g (Fig. 1). This is manifested in the lapse of 1-2 weeks between the spiking of the soil samples with the con taminan t s and the onset of an active biodegrada- t ion process. After the lag phase, mineral izat ion of each of the studied aminonaph tha lene compounds showed a two-phase process; an initial rapid phase followed by a second one with a diminished rate (Fig. 1). The diminished mineral izat ion rates, observed for the second phase, are a t t r ibuted to the partial unavailabil i ty of the con taminan t s to microbial uptake (Al-Bashir et al., 1994). Figure 1, also, shows that amount s mineralized (yg contami- nan t g ~ slurry) increased with increasing initial concentra t ion in slurry ( / lg of con taminan t g ~ soil slurry) but the latter decreased when expressed as a

20

16

12

0

m 16 ..=

0

16

12

l l l i

a. 1 - a m i n o n a p h t h a l e n e ~

~- ~ " / S i control x . / ~ - - +- S lag/g

~ . " " ~ 10 ~g/g . j ~ ----eP Z O ,ug/g

- , ~ f ' ~ ~ - - * - so.Q/g : ~ ~ ~ - - - e - 1 O0/~j/~ = r ~ : ~ - - % ~ , ~ s o ~g/,r

b. Z-aminonaphthalene

• control 5 ~g/g

~ 1 S ~g/g ~ 30/Jg/g

- ~ SO .ug/g 100/~g/g

• ~ . J . . ,'.t. x lSO /~g /g

c , 1 - amino -2 -me thy inaph tha lene

• control i I 0/Jglg

--'o--'20 jug/g + 35/~g/g

• SO/~/fl - ' - '~ 1 O0/.Jg/g

• = ' ~ . 200ua /a

0 50 IOO 150 ZOO Z50 Time (days)

Fig. 1. Mineralization of (a) I-aminonaphthalene, (b) 2- aminonaphthalene and (c) I-amino-2-methyl-naphthalene

at various initial concentrations in flooded soil.

Nitrogen-substituted naphthalenes in flooded soil--part I1 1829

O.Z

0.15

0.1

0.05

i O

0.15

I O d

• 0 . 0 5

0

0.15

, , ~ ' " = ,

, ,

Y T

i l i I i I i

0 . 0 5 - i m l n o - 2 ~

o 0 40 80 lZO 1SO ZOO

Initial concentration in slurry (/~/g)

Fig. 2. Initial mineralization rates as a function of total concentration in slurry of (a) l-aminonaphthalene, (b) 2-aminonaphthalene and (c) l-amino-2-methylnaphthalene.

percent of the total amounts added. Furthermore, the initial mineralization rates for the three aminonaph- thalenes (/Jg contaminant g - ' slurry day - ' ) increased with increasing initial total concentration in the slurry and reached their maxima at about 100/zgg -~ as shown in Fig. 2. In this figure and subsequent ones, an error bar indicates the standard deviation of three measurements. These errors are independent from each other as opposed to the dependent errors incorporated in the cumulative observations of Fig. 1 (Callas and Gehr, 1989).

Initial-phase mineralization

As Fig. 3 indicates the adsorption of the studied compounds gave rise to hyperbolic isotherms that were best described by the Langmuir model:

q = KLCM/(I + KLC ) (I)

where q is the sorbate concentration (gg g-~ soil) in the solid phase and c is solute concentration ( /~gml- ' ) in the liquid phase, KL is the Langmuir affinity parameters (ml/~g-~ ) and M is the adsorption maximal (#gml-J) . To find q in equation (1), the mass of contamination in the solid phase is calculated by subtracting the mass in aqueous phase from the total mass added initially and this in turn is divided by the weight of soil in the soil slurry. In Fig. 3, an

error bar represents the standard deviation of three measurements of the aqueous phase concentration.

Table 1 reports estimates of K L and M for the three aminonaphthalenes. These estimates were obtained using the Lineweaver-Burk method for the linear transformation of the Langmuir model (Kinniburgh, 1986). Accordingly, a plot of l/q versus 1/c yields a straight line such that KL=intercept/slope and M = l/intercept. Table 1 reports the r-values for the linear fits and the 90% confidence intervals for the estimated parameters. The confidence intervals of the two functions: (1) the inverse of slope and (2) the intercept over the slope, were obtained following the procedure outlined by Ang and Tang (1975).

When initial-phase mineralization rates are expressed as a function of contaminant aqueous- phase concentrations, they yield hyperbolic curves (Fig. 4) that fit Michaelis-Menten kinetics:

V = ( R m a x C ) / ( K m + C) ( 2 )

where, v, is mineralization rate (g g ml-J day-t) , c, is solute concentration (#g ml - ' ) in the liquid phase, Rm= is the maximum attainable rate (/zg ml- ' day- ' ) and K,, is the half-saturation constant ( /zgml- ' ) . Michaelis-Menton model assumes a reversible for- mation of a catalyst-substrate complex followed by a first order decomposition of the complex to a product. At relatively low substrate concentrations mineralization is first order with respect to the substrate concentration, i.e., v = (Rm,~/Km)c. While at very high concentrations, all of the catalyst is presumably complexed, and the mineralization rate thus reduces to v = R . . . . i.e., zero order kinetics with respect to substrate concentration (Fig. 4).

The linear transformation of the Michaelis-Menten equation using Lineweaver and Burk method gives (Piszkiewicz, 1977):

1/v = ( K m / R m a x C ) - 1 / R m a x ( 3 )

When l/v for each substrate, i.e., l-aminonaph- thalene, 2-aminonaphthalene, and I-amino-2-methyl- naphthalene, is plotted against its respective l/c for initial mineralization rates, a linear relationship is

700

600

500 o

.~ 400

300

200

1 O0 o o 0

0

r , . . . . ~ • ' + ' ' r • ' + ' ' '

~ n o - 2 - m e t h y l - naphthalene

, , i , i i , , i r , , , , , , , I , ,

10 ZO 30 40 SO SO 70 Concentration in liquid (/Jg/rnl)

Fig. 3. Partitioning of (a) l-aminonaphthalene, (b) 2- aminonaphthalene and (c) l-amino-2-methyinaphthalene between the solid and the aqueous-phases of the soil-water

suspension.

WR 2S/S--K

1830 BILAL AL-BASHIR et al.

Table 1. Estimation of adsorption and mineralization kinetic parameters of the studied amino-PAHs

Langmuir pa rame te rs * Michaelis-Menten First-order

kinetics parameterst kinetics parameters:~

K~ Rm~ . k,(day) I Compound K L M r (/~g/ml) (#g/ml.day) r x 10 4 r

I-aminonaphthalene 0.16_+0.06 765 + II0 0.998 4.0+2.1 0.42_+0.21 0.973 6,1 +_0.2 0,998 2-aminonaphtbalene 0,16 + 0.03 355 + 49 0.997 1.90 + 0.19 0.13 + 0.01 0,996 5.7 + 0.6 0.988 I-amino-2-methyl-naphthalene 0.33 + 0.15 1053 _+ 441 0,997 1.10 + 0.11 0.14 + 0.06 0,992 2.9 + 0.2 0,996

*K L is Langmuir partition coefficient (ml/#g), M is adsorption maximum (ug/g), r, is the correlation coefficient. "['Rma ~ is maximum mineralization rate attainable with increasing concentration [from equation (2)], K~ is the half-saturation constant [from equation (2)]. .~k. is the uptake coefficient from equation (3). The ( _+ ) error establishes the 90% confidence interval

obtained with correlation coefficients, r, reaching 0.973, 0.996 and 0.992, respectively. Table 1 sum- marizes the parameters Rma x and Km for all studied aminonaphthalenes and their corresponding 90% confidence intervals.

Not all of the compounds originally present in the aqueous phase were mineralized during the initial phase, especially those at concentrations leading to maximum mineralization rates. Pre- sumably, part of the soluble fraction of the aminonaphthalene underwent further irreversible adsorption (i.e., chemisorption) onto the soil, e.g., humic materials. To confirm this observation, the concentration of the contaminant in the aqueous phase was determined and was found to be negli- gible. Bollag et al. (1983), Parris (1980) and Hsu and Bartha (1974) have all reported that aromatic amines and their enzymatic metabolic products

.8

F: m

0.30 • 1 . i . i • i , i .

020 ? ] 0.10

0.00

0.20

0.10

0 .00

0.20

0.10

a. t-ami~naphthatene

, l = l , i i l l l =

b. 2.aminonaphthalene

I l l l l l i l l

c. 1-amino-2-methyl- naphthalene

0.00 I I I I I I I I

0 5 10 15 20 25 30 Concentrat ion in the aqueous-phase (/Jg/ml)

Fig. 4. Initial mineralization rates as a function of concen- tration in the aqueous-phase of (a) l-aminonaphthalene, (b) 2-aminonaphthalene and (c) l-amino-2-methylnaphthalene.

cross-link to humic substances in the soil through chemical bonding.

Second-phase mineral izat ion

The second-phase mineralization rates remained fairly constant during the experiment (Fig. 1), increasing linearly with the contaminant concen- tration in the soil part of the slurry and showing no sign of saturation (Fig. 5). The relationship between the mineralization rate and concentration in the solid phase can be expressed mathematically as:

v = k . q (4)

where ku is the substrate uptake coefficient from the solid-phase (day- ' ) and q is the substrate concen- tration in the solid-phase (#gg-~) . Equation (4) represents first order kinetics in which the rate is

"K

a 9

0.3

0.2

0.1

0.2

0.1

0.2

' i , i • i . i i

i i i i i , i i

b. 2<mlno~ap~halene

/ i i t i i

e. 1 -tm~no.2-molhy~lhllkm g

0 0 100 200 300 400 500 600

Concentrat ion in the solid-phase "q" ( pg /g )

Fig. 5. Second-phase mineralization rates as a function of concentration in the solid-phase of (a) l-aminonaphthalene, (b) 2-aminonaphthalene and (c) l-amino-2-methyl-

napthalene.

Nitrogen-substituted naphthalenes in flooded soil--part II 1831

directly proportional to the first power of substrate concentration in the solid phase. Plots of v against q for I-aminonaphthalene, 2-aminonaphthalene and 1-amino-2-methyl-naphthalene give straight-line relationships with correlation coefficients, r, reaching 0.998, 0.988 and 0.996, respectively. Values of k~ and their 90% confidence intervals are reported in Table 1.

General case: competitive adsorption

ks S A - - - ~ P + A + U /z[

S + A + U k6

S U

(5)

DISCUSSION

The biphasic curves observed in the mineralization of the aminonapbthalenes, i.e., l-aminonaphthalene, 2-aminonaphthalenes and l-amino-2-methylnaph- thalene, indicate a kinetic change. The initial phase of the mineralization curve is characterized by a readily available fraction of the contaminant which is in reversible equilibrium with the soluble substrate and which biodegrades following Michaelis-Menten reaction kinetics. Upon the depletion of the readily available fraction, the biodegradation process enters its second phase in which the desorption of the contaminant from the solid-phase becomes the rate-limiting step and the biological reaction is transformed into first-order kinetics.

In utilizing Michaelis-Menten kinetics to describe initial-phase mineralization rates, it is assumed that the catalyst concentration (the biomass concen- tration) remains constant with respect to the con- taminant concentration. This is a reasonable assumption given the fact that mineralization rates were measured at short and equal periods of time from the onset of the biodegradation experiments. Also the initial amounts of aminonaphthalenes con- stituted only a small fraction of the total organic matter available to the microbial population in the form of naturally occurring organic carbon and added methanol. Therefore, it is reasonable to ne- glect any difference in microbial growth prompted by variations in initial concentrations of the studied compounds.

In past kinetic studies, the presence of two different states of substrate bioavailability has been attributed to several reasons including, partitioning of the contaminant between soluble and sorbed forms (Steen et al., 1980, Mihelcic and Luthy, 1991), association of the substrate with active and inactive biological soil sites (Hamaker, 1972) and substrate partitioning between the soil labile and non-labile phases (Guerin and Boyd, 1990). In the present study, the catalyst suggested by the Michaelis- Menten model could refer either to microorganisms (enzymes) acting as the liquid/soil interface, to biologically active sites on soil surfaces or to labile matter associated with soil. Accordingly, the biologi- cal system under consideration can be described by reactions 5-7:

Initial-phase: Michaelis-Menten kinetics

k I kj S + A ~ SA -"~ P + A (6)

ks

Second-phase: first-order kinetics

k6 k3 SU--* SA --* P + A + U (7)

where S is a quantitative measure of the amount of the substrate, A represents the biologically-available sites which stand for the catalyst, U refers to the biologically-unavailable sites of the soil, SA and S U are the complexed forms of the substrate with available and unavailable sites, respectively, P is the reaction product and kl-k6 are reaction constants.

The overall reaction 5 describes the substrate, S, partitioning between the biologically-available and unavailable sites. The substrate undergoes competi- tive adsorption between these two sites, where k t/k 2 is a measure of the affinity of the substrate to the biologically-available phase and k4/k5 is a measure of the affinity of the substrate to the biologically- unavailable phase. Initially, when the substrate availability is not limiting which is a special case of reaction 5, the biological reaction is described by reaction 6 and follows Michaelis-Menten kinetics. However, when the substrate in the bioavailable fraction is depleted, the biological reaction is reduced to reaction 7, another special case of reaction 5. In this case, the rate of desorption of the substrate from the biologically-unavailable to the biologically- available phase constitutes the rate-limiting step. Here, the substrate becomes available to microorgan- isms as a result of desorption and the reaction proceeds only in the forward direction.

The previous discussion has dealt separately with each single compound and no attempt was made to investigate the effect of substitution on the mineraliz- ation kinetics. However, the following mathematical derivation seeks to do so. Under equilibrium conditions, the Michaelis-Menten reaction yields

k, [S][A ] = (k 2 + k3)[sa ] (8)

where [] indicate concentration of the respective species. Rearranging equation (8) yields

[S ][A ]/[SA ] = (k 2 + k 3)/kl = g m (9)

where K~, as previously defined in equation (2), is the Michaelis half-saturation constant.

According to equation (9), K~ represents the constant for the momentarily equilibrium (termed as

1832 BILAL AL-BASHIR et al.

pseudo-equilibrium) between the associated and dissociated forms of S and A. It is pseudo equilibrium because as mineralization proceeds, the equilibrium is continually disturbed.

Maximum reaction rate (Rma x) occurs when all the catalyst is present in the SA-complex form, in which case the maximum reaction rate can be expressed as

Rmax = k3[A ]total ( 1 0 )

where [A ],ot,~ is the total concentration of A that is available for complexing the substrate. Substitution equation (10) back into (9) yields

g m = ( k 2 q- Rmax/[A ]total)/kl (1 l )

Since the same soil type was used for the different compounds, then it is reasonable to assume that [A ],o,,t is constant across the three compounds. Also, it is important to emphasize here that while the available substrate is gradually being utilized by the microorganisms, a series of equilibrium states occur momentarily with their corresponding kj and k2- values [equation (6)]. However, it has been found that aminonaphthalenes exhibit significant hysteresis (see AI-Bashir et al., 1994), therefore, as mineralization proceeds, k z gradually becomes negligible while k~ remains constant. Given the above, then equation (11) can be reduced to

Km=CtRm,x/k j (12)

where • = I[A ]total- For an initial equilibrium state prior to the onset

of the mineralization process, the corresponding sorption equilibrium constant (K~q) is given by

Kcq = k I/k2. o (13)

where kz. 0 is the initial conditions desorption rate. From equation (13), we have

k, = flKeq (14)

where fl = k2. 0. Substitution equation (14) back into (12) yields

Km=Y(Rmax/Keq ) (15)

where 7 =~/fl. According to equation (15), Km is directly proportional to Rma x and is inversely proportional to the initial conditions equilibrium constant Koq. However, the initial equilibrium constant for adsorption was obtained for both avail- able and unavailable sites and it is reported in Table 1 as values of KL for the three aminonaphthalenes.

Using experimental data for the three aminonaph- thalenes compounds, a plot of Km against Rmax/K L gives a linear relationship with a high correlation coefficient, r equals to 0.994 (Fig. 6). Therefore, the plot suggests that KL and Km correlate and that in this specific case, KL is directly proportional to K~ for the biologically-available sites. In other words, the affinity to the biologically-available sites (k~/k2) rela- tive to the unavailable ones (k4/ks) remained constant across the three amino-compounds. However, the

oc

T I

2 3 4 K

m

Fig. 6. Relationship between Michaelis Menton mineraliz- ation constants and biological affinity parameters for the

three studied aminonaphthalenes.

correlation coefficient obtained in Fig. 6 is only indicative of a particular trend and does not fully reflect the accuracy of the estimated parameters.

In conclusion, for a better understanding of the biodegradation kinetics of recalcitrant organic contaminants in the soil environment, it is necessary to investigate the range of limiting factors influencing the biodegradation process. This will help distinguish between three kinds of recalcitrance: first, the inherent properties of the contaminant, i.e., k 3 is limiting, second, catalyst limitations, i.e., kj is limit- ing and finally the substrate bioavailability, i.e., k 6 is limiting. Recalcitrance of aminonaphthalenes in the soil environment is attributed to their ability to form complexes with soil matter which, to a large extent renders them biologically unavailable. Further work is needed to investigate the effect of soil organic content, cation exchange capacity and soil-to-water ratio on initial and second phase mineralization rates and total amounts mineralized. This has important implications for mediation of contaminated soil and ground water.

Acknowledgements--The authors greatly acknowledge Dr Attila Demeter and Dr Tibor Cseh for their help in synthe- sizing the radiolabetled compounds and Ms Chantale Beaulieu for her technical assistance.

This research paper is registered as NRC Number 33860.

REFERENCES

AI-Bashir B., Cseh T., Leduc R. and Samson R. (1990) Effect of soil/contaminant interactions on the biodegrada- tion of naphthalene in flooded soil under denitrifying conditions. Appl. Microbiol. Biotech. 34, 414-419.

AI-Bashir B., Hawari J., Leduc R. and Samson R. (1994) Behavior of nitrogen-substituted naphthalenes in flood soil--Part 1. Sorption/desorption and biodegradation. Wat. Res.

Ang A. and Tang W. (1975) Probability Concepts hi Engin- eering Planning and Design, Vol. I, Chap. 4. Wiley, NY.

Bollag J., Minard R. D: and Llu S. (1983) Cross-linkage between anilines and phenolic humus constituents. Envir. Sci. Technol. 17, 72-80.

Cailas M. D. and Gehr R. (1989) The oxygen uptake rate approach for analysing respirometric biochemical oxygen demand data--l. Method development. War. Res. 23, 985-992.

Nitrogen-substituted naphthalenes in flooded soil--part 1I 1833

Guerin W. F. and Boyd S. A. (1990) Influence of sorption/desorption on the bioavailability of organic con- taminants. Second year progress report to the Department of Energy, Washington, D.C.

Hamaker J. W. (1972) Decomposition: quantitative aspects. In Organic Chemicals in the Soil Environment, Vol. 1, Chap. 4 (Edited by Goring C., A. and Hamaker J. W.). Marcel Dekker, NY.

Hata K. (1971) New Hydrogenating Catalysts: Urushibara Catalysts. Halsted Press Div., Wiley, NY.

Hsu T. and Bartha R. (1974) Interaction of pesticides- derived chloroaniline residues with soil organic matter. Soil Sci. 116, 444-452.

Kinnibrugh D. G. (1986) General purpose adsorption isotherms. Envir. Sci. Technol. 20, 895-904.

Mihelcic J. R. and Luthy R. G. (1991) Sorption and microbial degradation of naphthalene in soil-water suspension under denitrification conditions. Envir. Sci. Technol. 25, 169-177.

Parris G. E. (i 980) Covalent binding of aromatic amines to humates--l. Reactions with carbonyls and quinones. Envir. Sci. Technol. 14, 1099-1106.

Piszkiewicz D. (1977) Kinetics of Chemical and Enzyme- Catalyzed Reactions, Chap. 5. Oxford Univ. Press, NY.

Scow K. M., Simkins S. and Alexander M. (1986) Kinetics of mineralization of organic compounds at low concen- trations in soil. Appl. envir. Microbiol. 51, 1028-1035.

Steen W. C., Paris D. F. and Baughmam G. L. (1980) Effect of sediment sorption on microbial degradation of toxic substances. In Contaminants and Sediments, Vol. 1, Chap. 23. Ann Arbor Science, Ann Arbor.

Thomas J. M., Yord J. R., Amador J. A. and Alexander M. (1986) Rates of dissolution and biodegradation of water- insoluble organic compounds. Appl. envir. Microbiol. 52, 290-296.

Vogel A. (1978) Textbook of Practical Organic Chemistry, 4th edition, Chap. IV. Longman Scientific & Technical, UK.