Estimation of nitrogen removal rate in aqueous phase based on δ15N in microorganisms in solid phase

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Estimation of nitrogen removal rate in aqueous phase based on d15N in microorganisms in solid phase Nobuhiro Kanazawa, Yoshikuni Urushigawa Department of Management Science and Engineering, Faculty of System Science and Technology, Akita Prefectural University, 84-4 Ebinokuchi, Tsuchiya, Yurihonjo, Akita 015-0055, Japan

art i cle info

ab st rac t

Article history:

Microbial nitrification and denitrification are important processes for removing nitrogen-

Received 7 February 2007

ous compounds in aqueous systems. Nitrogen removal rate estimation is essential for

Received in revised form

controlling nitrogen removal processes and modeling the nitrogen cycle in ecosystems. The

28 April 2007

model described the relationship between ammonium removal rate (aqueous phase) and

Accepted 26 May 2007

the nitrogen stable isotope ratio (d15N) of microorganisms (solid phase) when a coupled

Available online 3 June 2007

nitrification–denitrification process occurs and assimilation and advections are maintained

Keywords: Stable isotope Nitrification Denitrification Oxidation ditch Activated sludge

in a steady state. An oxidation ditch in a municipal wastewater treatment plant was evaluated for 3 years using the model. The ammonium removal rate was calculated from the data of d15N of the activated sludge, it correlated significantly with the observed removal rate. The isotope fractionation factor (e) was determined to be 5.5% by using a nonlinear method. The model and obtained factor value were applicable for standard activated-sludge processes performed in parallel in the oxidation ditch and a river watershed. The model may help illustrate nitrogen behavior in ecosystems. & 2007 Elsevier Ltd. All rights reserved.

1.

Introduction

The eutrophication of many ecosystems has led to an increased interest in nitrogen removal in the nitrogen cycle. Estimation of the nitrogen removal rate is important for controlling nitrogen removal processes in water treatment and for modeling the nitrogen cycle in ecosystems. A wellknown nitrogen removal process is biological sequential nitrification–denitrification that involves ammonium (NH+4 )  oxidation to nitrite (NO 2 ) and nitrate (NO3 ) followed by the   reduction of NO2 and NO3 to nitrogen gas (N2), as shown in Fig. 1 (details are omitted) (Ye and Thomas, 2001). Under  hypoxic conditions, NO 2 and NO3 derived from nitrification are often directly and immediately available for denitrification, following which NH+4 nitrification becomes rate limiting. This simultaneous reaction is called coupled nitrification–denitrification; it plays an important role in wastewater treatCorresponding author. Tel.: +81 184 27 2136; fax: +81 184 27 2189.

E-mail address: [email protected] (N. Kanazawa). 0043-1354/$ - see front matter & 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2007.05.051

ment and nitrogen transformation in sediments (Kremen et al., 2005). The majority of the estimates of nitrogen removal rates in such processes are based on the measurements of key compounds such as NH+4 , NO 3 , and nitrous oxide in the aqueous or gas phase. However, this process of estimation based on the measurement of water quality would make quantification of the removal rates difficult in situations in which, for instance, the inflow load varies or the route of water flow is complex. Hence, the development of an alternate method to estimate nitrogen removal rates will help in illustrating the behavior of nitrogen in the nitrogen cycle. The estimation of nitrogen removal rates using stable nitrogen isotopes has been studied. There are 2 ways of utilizing stable isotopes. One is a tracer method (Ashkenas et al., 2004) in which inorganic nitrogenous compounds or nutrients containing an enriched stable isotope are added at a

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Nomenclature

RA0, RA isotopic compositions of NH+4-N at initial condi-

C 0, C

Rstd, Rsmp isotopic compositions of standard materials

tions (f ¼ 1) and when nitrification proceeds

f F 0, F

NH+4 -N concentrations in aqueous phase at initial conditions and after reaction (mg l1) residual ratio of NH+4 -N in the aqueous phase inflow nitrogen load and nitrogen removal rate (tons N year1) removal rate of NH+4 -N (kg d1)

and samples

e isotope fractionation factor (%) d15NM0, d15NM nitrogen stable isotope ratio of sludge in solid phase at initial conditions (f ¼ 1) and at the time of sampling (%)

P Q inflow rate of wastewater (m3 d1) RM0, RM isotopic compositions (15N/14N) of microorganisms at initial conditions (f ¼ 1) and when nitrification proceeds

site at which nitrogen removal reactions are occurring, and the isotopic enrichment of the products of the reactions is measured. The other method measures changes in the natural abundance of stable isotopes instead of using a tracer (Yoshida et al., 1989; Brandes et al., 1998). The proportions of stable isotopes in biological materials vary due to physical, chemical, or biological processes. These changes are called isotope fractionations (Hoefs, 1997) and are indicative of the progression of reactions. In nitrification and denitrification reactions, heavier molecular species are enriched in the residual fractions of dissolved nitrogenous compounds in the aqueous phase as the reactions proceed (Handley and Raven, 1992; Barford et al., 1999; Casciotti et al., 2003; Lehmann et al., 2004). The relationship between the enrichment and reaction ratio can be analyzed by the Rayleigh equation (Mariotti et al., 1981) described later. Since the d15N of nutrients varies as the reactions proceed, the d15N in microorganisms assimilating the nutrients should also vary accordingly. The d15N of microorganisms records the cumulative information of the reaction in the aqueous phase, and it is expected to be useful for the estimation of the nitrogen removal rate in water. However, the d15N in the

15N

Nitrification

15N

Denitrification

15N

microorganisms involved in the nitrogen removal process may have been affected not only by the d15N in the nutrients but also by the isotope fractionations of assimilation, advection of microorganisms, and other factors. These effects have made the quantitative estimation of the nitrogen removal rate difficult (Voss et al., 1996), particularly in the natural environment, where the inflow load is hard to measure or control. Hence, in a controlled and well-maintained wastewater treatment plant, the mass balance could be regarded as a simple subdivision model of the natural environment. Therefore, the wastewater treatment plant is an appropriate subject to study the relationship between the isotope ratio of microorganisms and the nitrification and denitrification reactions. In this study, the d15N of the activated sludge and the water qualities at a municipal wastewater treatment plant that had a stable inflow load and nitrogen removal for 3 years were monitored. A model describing the relationship between the nitrogen removal rate and the d15N in the microorganisms in the solid phase was shown, and a tentative application of the model to natural environments was discussed.

2.

Experimental methods

2.1.

Samples

–

NO2

NH4+

15N

(15N ) Assimilation (15N )

N2 –

NO3

microorganisms Fig. 1 – Sequential nitrification–denitrification pathway.  Products of nitrification—NO 2 and NO3 —are reduced directly and immediately by coupled nitrification–denitrification. d15N of NH+4-N increases and that of N2 decreases as nitrification and denitrification  proceed; however, the d15N of NO 2 and NO3 will change only slightly by the counterbalance of isotope fractionations. Although the d15N of microorganisms decreases with assimilation, it increases with an increase in the d15N of NH+4-N when the assimilation rate is constant.

Samples were collected from a municipal wastewater treatment plant in the city of Honjo, Akita Prefecture, Japan. In the plant, Line A used an oxidation-ditch method and Line B used a standard activated-sludge method; the wastewater collected from the city had been treated by both the processes in parallel. Line A had a primary clarifier, an oxidation ditch, and a final clarifier. The oxidation ditch was elliptical, 4-m wide and 2-m deep with a circumference of 96 m, and was equipped with horizontal-axis mechanical aerators operated 24 h a day for circulation; no intermittent operation to increase denitrification (Furukawa et al., 1998) was reported. Line B had 4 bioreactors (an anaerobic tank and 3 aeration tanks) connected in series and a final clarifier. The wastewater flowed into the bioreactors directly without separation in a clarifier. The sizes of the bioreactors were 200, 300, 300, and 450 m3, respectively, with a common width of 6.0 m and a depth of 5.2 m. The inflow and outflow samples of the

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biological reactors were collected every 2 weeks for 3 years from April 2001 for Line A and for 10 months from July 2003 for Line B. The water samples were filtrated through a 0.2-mm membrane for analysis. Activated-sludge mixed liquors were centrifuged at 14,000 rpm for 1 min to collect sludge pellets. Each pellet was resuspended in pure water, centrifuged again at 14,000 rpm for 1 min, and frozen at 30 1C or below; it was then subjected to freeze drying to measure stable isotope ratios in the activated sludge.

2.2.

Analysis of water and activated sludge

The filtrated water samples in each sampling were measured for total organic carbon (TOC) concentration by a combustioninfrared absorption method (using a TOC-5000; Shimadzu Corporation, Kyoto, Japan), for NH+4 -N concentration by a coulometric titration method (MT-1; Central Kagaku Corpora tion, Tokyo, Japan), and for NO 3 -N and NO2 -N concentrations by the Griess Romijn method (UV1240; Shimadzu Corporation, Kyoto, Japan) with reagent kits (LR-NO2, LR-NO3; Kyoritsu Chemical-Check Laboratory, Tokyo, Japan). The random errors  in the analysis of NH+4 -N, NO 2 -N, and NO3 -N concentrations 1 were on average 0.2, 0.02, and 0.04 mg l , respectively. Flow rate, solid retention time, and hydraulic retention time were obtained from the process control data of the plant. The freeze-dried activated-sludge samples were measured for the stable isotope ratios of d13C and d15N using an elemental analyzer-stable isotope ratio mass spectrometer (EA-IRMS; EA1112-MAT253, Thermo Electron, Bremen, Germany). The isotope ratios of d13C and d15N were expressed by the ratio of the isotopic composition (R) (13C/12C or 15N/14N) between the sample (Rsmp) and the standard material (Rstd), i.e., the Vienna Peedee belemnite and the air, respectively, in per mil notation (%):   Rsmp  1  1000. (1) d15 N or d13 C ¼ Rstd d13C and d15N of the samples were determined by comparison with the reference gases carbon dioxide and N2, respectively, which were calibrated using the National Institute of Science and Technology standards. A histidine reagent was used as a standard during sample measurement, and the precision of the d13C and d15N measurements was 70.2% or higher.

3.

Results

3.1. Variations in water qualities and index of activated sludge in Line A The time-series variations of the concentrations of NH+4 -N,  NO 2 -N, and NO3 -N for 3 years in Line A are shown in Fig. 2(a). + The NH4 -N concentrations in the inflow and outflow waters of the oxidation ditch were 27.472.4 mg l1 (average7standard deviation) and 13.277.0 mg l1, respectively, and the outflow  concentrations of NO 2 -N and NO3 -N were low at 0.270.3 and 1 0.971.6 mg l , respectively. As all the inflow concentrations 1  or less (not shown), the of NO 2 -N and NO3 -N were 0.1 mg l  -N and NO -N were regarded as the products of detected NO 2 3

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the oxidation of NH+4 -N. The detected amounts of NO 2 -N and + -N were always lower than the NH -N removal amount, NO 3 4 indicating that the missing nitrogen was mainly removed by denitrification, and a high denitrification ratio, i.e., (1([NO 2+ + N]out+[NO 3 -N]out)/([NH4 -N]in[NH4 -N]out))  100, of 92%712% was maintained. Consequently, it was concluded that coupled nitrification–denitrification was occurring in the activated sludge. The values of d15N in the activated sludge and the NH+4 -N removal ratio, i.e., (1([NH+4 -N]out/[NH+4 -N]in))  100, varied from 0.6% to 7.9% (3.7%72.3%) and 52%725%, respectively, as shown in Fig. 2(b). The variation in the NH+4 N removal ratio was significantly correlated with that of d15N in the activated sludge (r ¼ 0.82, po0.001). The values of d13C in the activated sludge and the TOC removal ratio for dissolved organic matter were stable at 24.1%70.4% and 79%76%, respectively, as shown in Fig. 2(d), and the process appears to have maintained stable rates in the activated sludge. The solid retention time and hydraulic retention time were 9.772.1 d and 1571 h, respectively. The flow rate was initially 25007230 m3 d1, but was reduced to 7907190 m3 d1 after April 2003 due to a change in the plant operating conditions.

3.2. Variations in water qualities and the index of activated sludge in Line B In Line B, as shown in Fig. 2(c) and (e), the TOC removal ratio of the fourth reactor was stable at 77%75%, but that of NH+4 -N varied greatly from 0% to 85% (29728 mg l1). The accumula tion of NO 3 -N and NO2 -N in the effluent was low, and this was regarded as an indication of the coupled nitrification– denitrification conditions. The value of d13C in the activated sludge was stable at 24.7%70.1%, but that of d15N varied greatly from 1.3% to 5.2% (0.3%71.9%) (Fig. 2(c) and (e)). The variation in the NH+4 -N removal ratio was significantly correlated with that of d15N in the activated sludge (r ¼ 0.87, po0.001). The solid retention time and hydraulic retention time were 5.971.7 d and 1671 h, respectively.

4.

Discussion

4.1. Isotope fractionation of activated sludge under coupled nitrification–denitrification Evidence of nitrogen removal is indicated by the value of d15N in microorganisms; however, the value of d15N in microorganisms will be affected by its variations in nutrients resulting from nitrification and denitrification, isotope fractionation during assimilation, advection of microorganisms and solid nitrogenous compounds, and other factors, each of which would have a different effect on the d15N values in microorganisms. The effect of advection, i.e., isotopic dilution by the inflow of solid nitrogenous compounds having different d15N values compared with the microorganisms in the original reactor, will have a great impact on the measured values of d15N in microorganisms when larger volumes of advection occur rather than incremental growth of microorganisms. However, the influence of advection on the variation of d15N in microorganisms must be lesser at a site

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a Concentrations (mgN l-1)

40

: NH4+-N

in,

: NH4+-N

out,

: NO2- -N out,

:NO3- -N out

30 20 10 0

10

100

8

80

6

60

4 40

2 0

20

-2

0

10

100

8

80

NH4+ removal (%)

15 N (‰)

b

6

60

4 40

2 0

20

-2

0

NH4+ removal (%)

15 N (‰)

c

d 100 90 -23

80 70

-24

60 -25

50

-22

100

TOC removal (%)

13C (‰)

-22

90 -23

80 70

-24

60 -25

TOC removal (%)

13C (‰)

e

Apr-04

Jan-04

Oct-03

Jul-03

Apr-03

Jan-03

Oct-02

Jul-02

Apr-02

Jan-02

Oct-01

Jul-01

Apr-01

50

Date (month-year)

Fig. 2 – Time-series variations of water qualities in the oxidation ditch of Line A (a), d15N of activated sludge and NH+4-N removal ratio of Line A (b) and of Line B (c), and d13C of activated sludge and TOC removal ratio of Line A (d) and Line B (e).

where particulate matter is removed by sedimentation or filtration. Hence, there is low variation in the advection rate in a wastewater treatment process. In the nitrogen removal process under the condition of coupled nitrification–denitrification, residual NH+4 in water will be the dominant nitrogen source for microorganisms because the oxidative products, such as NO 3 , have been consumed by rapid denitrification. The isotope fractionation of residual NH+4 in water during

nitrification is expressed by the Rayleigh equation (Mariotti et al., 1981; Morasch et al., 2001; Hunkeler et al., 2001): RA ¼ f =1000 , RA0

(2)

where f is the residual fraction of NH+4 , e is the isotope fractionation factor, and subscripts A and 0 refer to NH+4 in water and the initial condition (f ¼ 1), respectively. Isotope

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fractionation during assimilation can be expressed as the ratio of isotopic compositions between microorganisms and residual NH+4 , and d15N in microorganisms must be smaller during assimilation (Fig. 1). However, if the growth rate of microorganisms is constant, the ratio of the isotopic compositions between microorganisms and residual NH+4 must be constant even if the isotopic composition of NH+4 assimilated by the microorganisms has varied. Therefore, the ratio of the isotopic compositions at the initial conditions between RM0 and RA0 is equal to the ratio of the isotopic compositions as nitrification proceeds between RM and RA as follows: RM0 RM ¼ , RA0 RA

(3)

where the subscript M refers to microorganisms. In general, the incremental growth of autotrophic ammonia-oxidizing bacteria can be ignored when compared with the overall incremental growth of the microorganisms (Donisi et al., 2002), and the growth rate of the microorganisms becomes independent from the progress of nitrification. The relationship between the residual ratio of NH+4 -N and the d15N of microorganisms under the conditions of coupled nitrification–denitrification and constant growth rate can therefore be expressed based on Eqs. (1)–(3) as d15 NM þ 1000 ¼ f =1000 . d15 NM0 þ 1000

(4)

Since ln(x+1)Ex when x51, Eq. (4) can be approximated as ! d15 NM  d15 NM0 f ¼ exp , (5)  where d15NM is d15N in microorganisms. Eq. (5) suggests that the residual ratio of NH+4 -N in water could be expressed as a function of d15NM in the solid phase. In the 2 treatment processes investigated in this study, the coupled nitrification–denitrification was considered to be in process, and substrate uptake into microorganisms had been maintained in a steady state, as described above. The parameters e and d15NM0 were then determined by Eq. (5) from the data on d15N in the activated sludge and the NH+4 residual ratio for 3 years in Line A. The parameters e and d15NM0 were determined by a nonlinear method of trial and error. The optimal solution of the parameters was found by using the solver add-in Microsoft-Excels until the sum of the square error between the calculated f and the observed f was minimized; the values were determined to be e ¼ 5.5% and d15NM0 ¼ 1.1%. In the calculation, it was regarded that NH+4 consumption for the growth of the microorganisms was almost equal to NH+4 production by the degradation of suspended solids in wastewater; the NH+4 residual ratio f was defined as (1[NH+4 -N]out/[NH+4 -N]in). Although an excess activated sludge of 75 mg l1 (6.0 mg N l1) was produced at the plant on average, several data suggested that approximately just 2 mg l1 of NH+4 -N (Fig. 2(a)) appears to have been removed. If all the excess activated sludge was generated from dissolved NH+4 -N, at least 4 mg l1 of fresh NH+4 -N must be produced from the suspended solids in the oxidation ditch. If the NH+4 consumption for growth was greater than its fresh production from suspended solids by 2 mg N l–1, the parameters were e ¼ 6.6% and d15NM0 ¼ 1.1%. However, the

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excess sludge includes not only mature microorganisms that consumed the dissolved NH+4 but also unreacted suspended solids that are included in the influent. Although the proportion of microorganisms to unreacted solids in excess sludge was not clear, the difference between the consumption and production of NH+4 was estimated to be lesser than 2 mg N l–1. Therefore, the isotope fractionation factor (e ¼ 6.6%) was overestimated and then it was regarded that the NH+4 consumption for the growth was almost equal to the NH+4 production as described previously. Although variations in d15N in activated sludge were also correlated with the nitrogen removal ratio as defined by ([NH+4 -N]out+[NO–2-N]out+[NO–3-N]out)/[NH+4 -N]in (r ¼ 0.78, p ¼ 0.014), it was not correlated with the denitrification ratio (1([NO–2-N]out+[NO–3-N]out)/([NH+4 -N]in[NH+4 -N]out)) (r ¼ 0.09,  p ¼ 0.81). These results indicate that NO 2 and NO3 do not affect d15N in activated sludge. This absence of effect is considered to be the result of poor assimilation or due to the  + similarity of the d15N of NO 2 and NO3 to that of NH4 due to the counterbalance of the isotope fractionations of nitrification and denitrification (Fig. 1). The latter reason may be more  likely because NO 2 and NO3 are intermittently produced in the coupled nitrification–denitrification process, and the experiments with Nitrosomonas europaea showed that the isotope fractionations during both nitrification and denitrification were almost the same at eE35% (Mariotti et al., 1981; Yoshida, 1988). The parameter e for d15N of activated sludge or microorganisms versus NH+4 concentration under coupled nitrification–denitrification has not been reported; therefore, a comparison with the parameters for the other site under the same condition could not be done. However, under the condition of NO 3 consumption in marine environments, e for d15N of plankton versus NO 3 concentration was reported to range from 2.5% to 5% (Altabet and Francois, 1994;  Wada, 1980) and that for d15N of NO 3 versus NO2 concentration was 3% to 9% (Voss et al., 1996; Sigman et al., 1999, 2001). It is interesting that the values are within the same range as those obtained in this study. The mechanism for the increase of d15N in microorganisms under the condition of NO 3 consumption can be considered to be almost similar to that proposed for the model in this study if it is explained in terms of the assimilation of the enriched NO 3 during the denitrification process. The relationship between the d15N of microorganisms and the NO 3 removal rate will then be analyzed using the same model. However, since the growth rate of microorganisms under anaerobic conditions for denitrification is generally slow, the microorganisms that had an increased d15N during the denitrification process should have had enough time to assimilate the nutrients. The advantages of the nonlinear method for determining e and d15NM0 are described below because the description regarding a problem in the logarithmic simple linear regression analysis (the linear method) has not been reported. Although the linear method for calculating the relationship between d15NM and ln f analytically in order to obtain e and d15NM0 is often used in studies (e.g. Mariotti et al., 1981; Barth et al., 2002; Lehmann et al., 2003), the nonlinear method was deliberately used instead in this study. As shown in Fig. 3, the calculated estimate by the linear method (dashed line, r ¼ 0.77, versus actual values) was worse than the result

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obtained by the nonlinear method (solid line, r ¼ 0.81), and the linear method provided different results (e ¼ 2.9% and d15NM0 ¼ +1.0%) from those obtained using the nonlinear method (e ¼ 5.5% and d15NM0 ¼ 1.1%), which was a serious problem. The linear method could not provide all the real optimal solutions in the analysis with its logarithmic system. The smaller ln f tended to have a larger random error or lower accuracy, whereas overvaluation at smaller ln f and undervaluation at larger ln f occurred in the linear method. This phenomenon is a mathematically inevitable problem due to the use of logarithms. The nonlinear method is therefore recommended over the linear method since the linear method does not always provide an optimal solution of the isotope fractionation factor.

1.0 Nonlinear Linear

0.8

f

0.6 0.4 0.2 0.0 -2

2

0

4

6

8

10

15N (‰) Fig. 3 – Comparison of linear and nonlinear fitting methods for determination of the isotope fractionation parameters.

4.2. Application of d15N of activated sludge for estimation of ammonium removal rate Since the NH+4 residual ratio could be obtained from d15N in activated sludge by Eq. (5), the NH+4 removal rate could also be derived when an inflow load was given. As the inflow load was expressed as C0  Q, where C0 is the NH+4 -N inflow concentration (mg l1) and Q is the inflow water quantity per day (m3 d1), the NH+4 -N removal rate P (g d1) was expressed as P ¼ (1f)  C0  Q  103 or " !# d15 NM  d15 NM0 (6) P ¼ 1  exp C0 Q  103 .  Then, the average of the inflow NH+4 -N concentration and flow rate, d15N in the activated sludge in every sampling in Line A, and the parameters e ¼ 5.5% and d15NM0 ¼ 1.1% were substituted into Eq. (6), and the NH+4 -N removal rate was estimated. Comparisons of the calculated removal rates with the observed values are shown in Fig. 4(a, squares and circles); the correlation coefficient was r ¼ 0.87 (po0.001), which was significant, and the standard error was 6.0 kg d1. As the residual ratio can be estimated from data on the isotope ratio, the inflow concentration can also be calculated by C0 ¼ C/f, where C is the NH+4 -N concentration after reaction (mg l1). Then, Eq. (6) may be rearranged as " ! # d15 NM0  d15 NM P ¼ exp  1 C Q  103 . (7)  This allows the use of NH+4 -N concentration data at a site where a sample of microorganisms is collected for use in the estimation of the NH+4 removal rate. Although the estimation

NH4+ -N removal (kg•d-1)

a 90 Observation Estimation by eq. (6) Estimation by eq. (7)

60

30

0 90 Observation Estimation

60 30

Apr-04

Jan-04

Oct-03

Jul-03

Apr-03

Jan-03

Oct-02

Jul-02

Apr-02

Jan-02

Oct-01

Jul-01

0 Apr-01

NH4+ -N removal (kg•d-1)

b

Date (month-year) Fig. 4 – Comparison of NH+4-N removal rate between observed values, estimated values from the d15N of the activated sludge with average of inflow NH+4-N concentration in Line A (a, circles) and in Line B (b, circles), and the estimated values from the d15N of activated sludge with each outflow NH+4-N concentration in Line A (a, triangles).

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at the site where the inflow load is not known cannot be achieved with only the concentrations of nitrogenous compounds in the site, which is often the case in natural environments, Eq. (7) enables the estimation of the NH+4 removal rate even in such a case. In the case of Line A, when the data on the outflow NH+4 -N concentration was applied to Eq. (7), the result of the estimation was significantly correlated (r ¼ 0.63, po0.001) and the standard error was 12.7 kg d1 (Fig. 4(a, triangles)), but the result was inferior to that obtained by using Eq. (6). The NH+4 -N concentrations were the values of the samples collected every 2 weeks, and the hydraulic retention time (15 h) was shorter than the sampling interval and solid retention time (9.7 d). The d15N value of the activated sludge might reflect the cumulative information of the activated sludge remaining in the reactor, whereas the NH+4 -N concentration was taken as an instantaneous value. Therefore, it is reasonable that there are small differences among the estimated and observed NH+4 removal rates described above. In order to determine the applicability of the model and its parameters, the NH+4 removal rate in Line B was also estimated with the same procedure in Eq. (6) using the parameters e ¼ 5.5% and d15NM0 ¼ 1.1% obtained from Line A and the other values observed in Line B. The correlation between the variation in the estimated values and those of the observed ones was significant (r ¼ 0.90, po0.001) and the standard error was 4.8 kg d1, as shown in Fig. 4(b). The error bar in Fig. 4 shows the error of estimation derived from the measurement error of d15N. As discussed above, it was concluded that the NH+4 removal rate could be estimated by Eqs. (6) or (7) with the data on the inflow or outflow of the NH+4 concentration, flow rate, and d15N in the activated sludge, suggesting that the model and the isotope fractionation parameters had some generality.

4.3. Tentative application of the model to natural environments In this study, the range of d15N in activated sludge (0.6% to 7.9%) was similar to that observed in sediments of the Yodo River watershed (0.9% to 7.8%) reported by Yamada et al. (1996). Lake Biwa located in the Yodo River watershed can be regarded as a reactor where the denitrification of NO 3 is occurring. The retention time of water is 5.5 years (Kita et al., 2006); this is sufficient time for the growth of microorganisms with NO 3 assimilation even under anaerobic conditions, and, therefore, the same analysis as that used in this study was applied. The NO 3 removal rate in Lake Biwa was estimated using the data reported by Yamada et al. (1996): F0 ¼ 4.0  103 tons N year1, d15NM0 ¼ 2.0%, and d15NM ¼ 6.2%, where F0 is the inflow load (tons N year1). Eq. (5) could be rearranged as    dM  dM0  F0 , (8) Fd ¼ 1  exp  where Fd is the denitrification rate (tons N year1). Since the isotope fractionation factor for this site was not determined, the supplied values and our determined parameter e ¼ 5.5% were applied to Eq. (8). The result was Fd ¼ 2.1  103 tons N year1, which agreed with the results obtained by Miyajima

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(1994) and Yamada et al. (1996); half of the nitrogen was removed. In fact, NO 3 removal in a lake does not involve only the denitrification reaction. Planktons assimilate NO 3 s at the surface layer and sink to the lake bottom; further, the coupled nitrification–denitrification process must occur at the surface of sediments. However, the variation in d15N in sediments will inevitably be linked to each step of the reactions if the reactions are in a steady state. This is because the isotope fractionation of each reaction is maintained constant, and the d15N in sediments is determined by the integration of the isotope fractionations in each reaction. In the food web, it was found that d15N in an organism increases from 3% to 5% with increasing trophic levels (Minagawa and Wada, 1984). This is a case in point that although food webs involve several steps, the overall variation of the isotope ratio is maintained constant in steady-state ecosystems. In the case of the wastewater treatment system, in this study, in which it was assumed that assimilation and advection were controlled in a steady state, the coupled nitrification–denitrification process was predominantly responsible for the variation in d15N in activated sludge; this relationship was expressed as the model, and an isotope fractionation factor could be determined as described above. In other words, excluding a predominant process responsible for the variation of d15N in the solid phase, the reactions in the nitrogen removal process are regarded to be in steady state. Then, this model can be used for the estimation of the nitrogen removal rate even in a natural environment. However, further confirmation of the isotope fractionation factor between nitrogenous ions in the aqueous phase and microorganisms in the solid phase is needed before it can be applied for the process and more experiments at various sites are required. The concept of the model and the isotope fractionation factor used in this study may have broader applicability and are expected to be utilized to illustrate nitrogen behavior in ecosystems.

5.

Conclusions

This model theoretically described the relationship between the d15N of microorganisms in the solid phase and the NH+4 removal rate in the aqueous phase. This model is based on the assumption that coupled nitrification–denitrification occurs, and the growth rate of the microorganisms is constant. Activated sludge in a municipal wastewater treatment plant was monitored for 3 years, and the following conclusions were drawn by application of the observed data to the model:

(1) The model was found to be acceptable in view of the data from the oxidation ditch in a municipal wastewater treatment plant. The parameters indicative of the isotope fractionation were e ¼ 5.5% and d15NM0 ¼ 1.1%. (2) The model and its parameters were found to be applicable to other wastewater treatment systems and to a river watershed. Thus, the analysis of d15N in the solid phase could be a valuable tool for illustrating nitrogen behavior in ecosystems.

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Acknowledgments The authors wish to thank Mr. Toru Kagaya, Mr. Toshiyuki Sasaki, and Mr. Tomio Abe of Tohoku Kankyo Kanri Co., Ltd., for their cooperation in obtaining samples, the city of Yurihonjo for assistance with our research at their wastewater treatment plant, and Ms. Makiko Endo and Mr. Syusei Saito for sample measurements. R E F E R E N C E S

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