-------
            CHESAPEAKE  BAY
   PATUXENT TRIBUTARY
              r    i    i    i    i
                                        0
                                        2
                                        4
                                        6

                                        8
                                       10

                                       12
           i	  i    I     j
            -100  0   100  200  300
     -100 0   100  200  300
          POTOMAC TRIBUTARY
   CHOPTANK TRIBUTARY
    e
    o
   Q.
   U  ft
   0  8

       I 0

       12
              I    I    I     I    I
 2

 4

 6

 8

I 0
12
           -100   0   IOO  200 300          -100  0   IOO 200  300    j
                         REDOX  POT ENTI AL , m V
Figure 2.  Redox potentials (based on  platinum electrode) of estuarine sedi
          ments at  ten stations in Chesapeake Bay.

-------
                 >»
                 •o
                 c
         Q.
         O
         £.

         O
      o
      ID
   o
   o
                               o
                               m
                               
-------
at nitrate concentrations from 10 to 50 uM,  however,  at  nitrate  concentra-
tions > 100 ^ there was nearly a linear increase in  N20 concentration
(Fig. 4).  All other stations exhibited saturation in N20 production  at
nitrate concentrations <_ 150 yM and most at  much  lower concentrations.

     In general,  there was a first-order relation between denitrification
potentials and NO^ concentration at relatively  low NO^ and zero-order
kinetics at higher NO^ levels (Fig. 5).  A rectangular hyperbolic  model
fit by the linear transformation of Dowd and Riggs (1965) adequately
described these relationships for all  but  two stations (table  4).   Correla-
tion coefficients (r^) for this Michaelis-Menten  type model  ranged  from  0.73
to 0.98 for eight of the stations, with only Sta. 3 and  4 having non-
significant relationships (r^ = 0.04 and 0.30,  respectively; Table  4).
Table 4.  Michealis-Menton kinetic constants including maximum denitrifica-
          tion and Ks  (half-saturation constant)  for estuarine sediments  in
          Chesapeake Bay.
STATION


1
2
3
4

5
6

7
8

9
10
MAXIMUM
nmol N'cc

6.99
6.67
*
*

2.65
124.35

57.14
24.37

77.33
222.30
DENITRIFICATION
•h nmolN-gdw" h
CHESAPEAKE BAY
9.8
24.2
*
*
POTOMAC TRIBUTARY
8.0
401.3
PATUXENT TRIBUTARY
146.4
62.9
CHOPTANK TRIBUTARY
151.5
555.6
Ks r2
(M

8.4 0.85
8.5 0.87
* 0.04
* 0.30

2.0 0.92
67.2 0.89

22.6 0.94
78.2 0.73

10.6 0.98
92.2 0.76
*Linear transformation models were nonsignificant
-------
     •o
     o»
          40
     ±    30

     e
     c
          20
z
o
\-

o

£    10
    LL)
    O
                      50      100     150     200    250
                               NITRATE,
Figure 4.  Denitrification potentials  (based on  N20 production) at different

          concentrations of nitrate at Sta. 3  in Chesapeake Bay.
-------
            o
            6
           Z
           o
           O
              |50
              100
           o
           -  50
               1.0
           77,0.5
           O
                    B
                          50
                                        mox = 151.5
                                               10 .6
                                                          1
        100
150
200
250
                                         Y = (0.0066) X + 0 07

                                                      =  151 5
                mo*  0.0066
                                                0.07  = 10.6
                                               0.0066

                                         r   = 0.98

                                          I	I
50      100      150

         [NO,] , /i
                                                 200
                250
Figure 5.   A)  Response of denitrification potentials at different concen-
           trations of nitrate in estuarine sediments at Horn Point.  B)
           Linear transformation of data in A)  which was used to calculate
           Michaelis-Menton kinetic constants of denitrification rates.
-------
     Nitrate-saturated denitri fication rates (9max) ranged from 8.0 to 555.6
nmol N gdw~l h~  , with the highest rates occurring in sediments located in
the oligohaline  regions of the Potomac and Choptank River estuaries (Table
4)«  Dmax was higher in sediments of the Choptank than Patuxent River
estuary, and rates in tributary stations tended to be higher than those in
the mainstem Bay.  Where kinetics conformed to the Michaelis-Menten model,
half-saturation  constants (Ks) ranged from 2.0 to 92.2 yM N0§ (Table 4).
The tributary stations had Ks values > 60  ^ in the oligohaline regions
and < 25 yM in mesohaline regions.  The only two mainstem Bay stations
that had significant linear models (Sta. 1 and 2) had similar Ks values that
were about 8.5 jM (Table 4).

                             Regressi on _Mode_l_s_

     Significant regression models were found for both independent variables
KS and Dmax, using stepwise regression analysis with selected sediment
characteristics  (Table b).  Independent variables included only those
observations from stations that had significant kinetic constants, thus
excluding Sta. 3 arid 4.  Only the N:P ratio (x) of surface sediments was
significantly related to KS (Y), where (Y = -8.394 (X) + 120.136, r  = 0.71)
(Table 5).  Dmax(Y) was correlated with two independent variables including
•salinity (XI) or bottom water and TC (X2) of surface sediments  (Y = -34.22
(XI) + 8.06(X2)  + 348.27, r2 = 0.79) (Table 5).  Both models were significant
at P levels < 0.025.
Table 5.  Models of factors that correlate with Michaelis-Menton constants
          Dmax  (maximum denitrification potentials) and Ks (half-saturation
          constants) based on stepwise regression analysis.
Dependent
Variable
KS
^max
Independent
Variables
N:P Ratio
Intercept
Salinity
Carbon
Intercept
Slope r2
-8.394 0.705
120.136
-34.22 0.790
88.06
348.27
P. > F
0.0092
0.0205
-------
                                 DISCUSSION
      It is not apparent why denitrification potentials in two of the main-
stem  stations had no kinetic relation to increasing NO^ concentrations,
since the levels of electron source  (organic carbon) at these stations were
similar to those in the upper tributaries  (Table 4).  However, these two
stations are located in a region of  the Chesapeake Bay where the overlying
water is commonly anoxic during the  summer  (8oynton et al. 1985), and
associated high hydrogen sulfide concentrations occur at the surface of
these sediments (e.g. 13.8 M; Hill 1984).   Concentrations of sulfide at half
the reported concentrations in these sediments will prevent the inhibitory
action of acetylene on ^0 production thereby causing an underestimate of
.denitrification by this method (Tarn  and Knowles 1979).  On the other hand,
Stfrensen et al . (1980) have also observed  that high sulfide concentrations
block the formation of N^ from NO or ^0,  thus possibly counteracting this
methodological artifact in sulfide rich sediments.

      These low denitrification rates in Sta. 3 and 4 may also be associated
with  the observation that NO^ reducing bacteria occurring in sediments
rich  in Fe   ions and low in NO^ concentration can utilize iron rather
than  nitrate as an electron acceptor under  conditions of high organic carbon
content (Stfrensen 1982).  Total iron concentration sof sediments in this
region of Chesapeake Bay are relatively high with values often exceeding 4%
dry wt (Helz 1985), and during October, NO^ concentrations in the over-
lying water are generally < 5  \]fl.  With the high concentration of organics
in these sediments, this evidence strongly  suggests the possible utilization
of Fe   by N0§ reducing bacteria.  The stimulation of denitrification in
the sediments at Sta. 3 by NOj concentrations > 100 jM, is consistent
with  the findings of Stfrensen (1982), who  showed that the iron reduction by
denitrifiers could be superceded by  amendments with 200  (M of NO^.  At
this  stage, of course, these explanations  for low denitrification potentials
of Sta. 3 and 4 are only speculative, and  other effects such as various
metals (Slater and Capone 1984), which are  known to exist at relatively high
concentrations in these Bay sediments (Helz 1982).

      The response of denitrification potentials to NO^ concentrations
followed rectangular hyperbolic type kinetics in the other 8 stations of
Chesapeake Bay sediments investigated.  These kinetic relations could be
characterized by the constants Dmax  anc' Ks  tna*- are the maximum denitrifica-
tion  rate and half-saturation constant, respectively.  Reporting zero-order
response of dentrification to NOg concentration, Kaspar (1982) argued that
the kinetic nature of denitrification observed by Oren and Blackburn (1979)
was possibly due to diffusion.  The  well-mixed slurries with high water:
sediment ratios used in our experiments exclude diffusion as a rate-limiting
process.  Our results demonstrate that in  most cases denitrification
-------
potentials in Chesapeake Bay sediments will  respond to increasing N0§
concentrations in a kinetic fashion.

     This kinetic response and the constants used to describe it resemble
those used to interpret first-order enzyme kinetics (Dowd and Riggs 1965).
However, eventhough the curves may be similar, the interpretation of these
constants for enzyme systems may not be applicable to complex heterogenous
populations (Williams 1973; Goldman and Glibert 1983).  In our studies it
may be assumed that Ks represents a measure  of the physiological affinity of
the denitrifier community for NOj substrate.  Oenitrification potentials
were not normalized to microbial biomass, so that Dmax may be taken as an
index of the density of denitrifying bacteria in a constant volume of sedi-
ment (McLaren 1976).  However, interpretations of Ks values should not be
affected by bacterial abundance as long as no physiological adaptation of
the population occurs during the course of the experiment  (Nedwell 197b).

     The range in Ks values observed for Chesapeake Bay sediments is similar
to that for measurements in other estuarine systems (Table 6).  KS values <
25  iM, which were common in mainstem Chesapeake Bay have been reported for
Japanese estuarine systems (Koike et al. 1972; Koike et al. 1978; Koike and
.Hattori 1979) and for sediments in the Bering Sea (Koike and Hattori 1979).
Average values observed for Patuxent and Choptank River sediments (about
50  jM) are similar to those measured for the Scheldt estuary in Belgium
(Billen 1978) and Izembek Lagoon, USA (lizumi et al. 1980).  One excep-
tionally high Kr value (344 jM) has been reported for Kysing Fjord (Oren
and Blackburn 1979).  Other sediments with much higher nitrogen loading
characteristics such as those adjacent to sewage effluents or ditches
draining agricultural fields have corresponding higher Ks values (300-250  jM
and higher denitrification rates (Nedwell 1975; Van Kessel 1977; Kohl et al.
1976).  However, these extreme K  values are based on NO^ uptake by intact
                                it effects of diffusion rather than enzyme
sediments, and may thus represent
Table 6.  Half-saturation constants (Ks) describing the kinetic response of
          denitrification (or nitrate reduction) rates to nitrate concentra-
          tions for marine and freshwater sediments, and soil.
      Site
                                         Process
Reference
Chesapeake
Bay, USA
Bering Sea
Tokyo
Odawa
Bay,
Japan
Estuary
Mangoku-Ura
Soldi
er Key
, Japan
, USA
2.0 - 92.2 Deni
3.5 - 9.1 Deni
24 Deni
25 Deni
27 - 42 Deni
30 Deni
tri
tri
tri
tri
tri
tri
fi
fi
fi
fi
fi
fi
cation
cation
cation
cation
cation
cation
Thi
Koi
Koi
Koi
Koi
s study
ke
ke
ke
ke
Capone
and Hattori 1979
et
et
et
al
al
al
and
. 1978
. 1972
. 1978
Taylor 1980
-------
Table 6.  Continued
      Site
                             Process
Reference
Scheldt Estuary, Belgium    50

Izembek Lagoon, USA         b3
(eelgrass sediments)

Kysing Fjord, Denmark      344

Vatuwaga River, Fiji    180 -   600
(mangrove sediments)

Danish Lakes
   Aerobic Sites
   Anaerobic Sites

Ditch Bank Sediments

Arable Soi1

Agriculture Soi1
                          Denitrification   Billen 1978

                          Denitrification   lizumi  et  al.  1980


                          Denitrification   Oren  and Blackburn 1979

                          Nitrate  Uptake   Nedwell  197b
376 -
7 -
14
4
290 -
6,027
893
,696
,759
3,479
Ni
Ni
Ni
Ni
trate
trate
trate
trate
Uptake
Uptake
Uptake
Uptake
Anderson 1977
Van
Van
Koh'
Kessel
Kessel
1 et al
1977
1977
. 1976
kinetics.  These methods may also overestimate actual  denitrification poten-
tials since a portion of the nitrate may be reduced^to end products other
than those
1975).
associated with denitrification (e.g.  NH^;  Stanford  et  al
     In Chesapeake Bay, higher Ks values occurred in the landward stations
of the tributaries where springtime NO^ concentrations are much higher
than the downstream stations (Kemp and Boynton 1984; Ward and Twilley 1986).
This pattern suggests higher KS values may represent enzymatic adaptations
to higher substrate concentrations (Fenchel  and Blackburn 1979).   Yoshinari
et al. (1977) observed that amendments of soils with glucose caused an
increase in KS for denitrification, suggesting that the generally higher
organic deposition rates in these estuarine regions (Kemp and Boynton 1984;
Boynton and Kemp 1985) may partially account for this pattern as  well.

     Nitrate-saturated denitrification potentials were generally  higher for
sediments in the Choptank than for those in the Patuxent, and rates were
generally higher in the tributary stations than in the mainstem Bay.  Simi-
lar spatial patterns among Choptank, Patuxent and mainstem Bay sediments
have also been observed for ambient and potential dentrification  rates using
l^N-NOg (Twilley and Kemp, Unpublished; see Fig. 6).  Springtime  nitrate
concentrations in the oligohaline zone are generally > 100 \p\ in  the Chop-
tank (Ward and Twilley 1986), compared to 50-75  \ft in the Patuxent (Kemp
and Boynton 1984), and 25-50 uM in the mainstem Bay (Boynton and  Kemp
-------
               30
               20
                10
           Kl
            E
            o
0
            o  20
            o
            H  10
            O
            u.
            £E
            -   0

            o  20
                0
                0
           CHOPTANK  RIVER
                           PATUXENT RIVER
                           CHESAPEAKE   BAY
                    TIDAL  FRESH  BRACKISH  MESOHALINE
                       (0-2%o)    (4-8%o)     (6-!2%o)
Figure 6.  Denitrification potentials of estuarine sediments based on amend-
          ments of 100 pM of 15N labelled  nitrate (Twilley and Kemp,
          unpublished data).
-------
1985).  Both Koike et al . (1972) and Stfrensen (1982)  have suggested that
higher concentrations of NO 3 may stimulate the enzyme activation  level  of
NO^ reducing bacteria.  Thus, while the affinity for  NO^ (K )  by  denitri-
fiers may be lower in regions with higher NO^ concentration,  increased
denitrification capacity (Umax), which reflects differences in population
density, may also result from higher ambient NO^ concentrations.

     Ambient denitrification rates, estimated from these kinetic  relations
(Table 4) in conjunction with ambient NO^ concentrations in overlying water
(Table 2), ranged from 1.4 to 123.3 nmol  N-gdw  'h'1  in Chesapeake Bay  in
October.  These rates are only applicable for the top few mm of sediment, and
rates would decrease as NU^ levels in pore waters decline with depth.  Most
denitrification rates reported for other estuarine sediments  are  < 1U nmol
N'gdw" 'h~  (Table 7) except for those amended with NO^ (Koike and Hattori
1978a).  Hattori (1983) calculated denitrification rates for eutrophic  Tama
Estuary [150 nmol N'gdw'^'h""   from Nishio et al. (1981)] that are similar
to the upper range in values for Chesapeake Bay.  Combining the kinetic con-
stants from our study with higher NO^ concentrations  (ca. 100   ^l) that
occur in the Choptank and Patuxent River estuaries during the  spring, we
estimate that denitrification rates at our stations range from 26 to 289 nmol
N'gdw'l'h"* in spring.  These estimated ranges suggest that actual  denitrifi-
cation rates may vary as much spatially due to the kinetic nature of denitri-
fication capacity as seasonally due to differences in NO^ concentration.

     A significant portion of the variability observed for Dmax and Ks  in
Chesapeake Bay sediments can be explained statistically in terms  of gross
sediment characteristics.  Nitrate-saturated denitrification  rates (Dmax)
were inversely correlated with salinity (r^ = 0.65).   Elsewhere,  it has been
shown that mean NO^ concentration in overlying water  is also  inversely
related to salinity in Chesapeake Bay (Kemp and Boynton 1984;  ward and
Twilley 1986).  Thus, salinity serves as a good predictor of  long-term  mean
NO-j levels.  As we have discussed above,  higher ambient NO^J concentra-
tions presumably lead to higher growth rates and more abundant denitrifier
populations.  Dmax was also directly correlated with  TC in sediments, which
is consistent with the concept that increased availability of  electron
donors can also stimulate denitrifier growth (Hattori 1983).   Other investi-
gators have demonstrated higher denitrification capacities with higher
organic matter concentrations in sediments (Terry and Nelson 1975; Koike and
Hattori 1978b) and soils (Bowman and Focht 1974; Yoshinari et  al. 1977).

     Similarly, Ks was inversely correlated to salinity (although the rela-
tion is less significant than for Dmax);  again, this  suggests  a lower
affinity for NO^ for those denitrifier populations generally exposed to
higher substrate concentrations.  K$ was also significantly related to  N:P
ratio of sediment particulates.  This unexpected result might  reflect the
effect of higher denitrification rates decreasing the residual particulate
nitrogen levels in these sediments.  Thus, it appears that sediment charac-
teristics influence the kinetic relations for denitrifier bacteria in
estuarine sediments in various ways.  The correlative information presented
in this paper suggests that physical and chemical features of  the sediment
and overlying water are largely responsible for the observed  ranges in
kinetic parameters for sediment denitrification in this estuarine system.
-------
Table 7.  Denitri fi cation rates  (nmol N-gdw"-*- •h"-'-)  of  marine and freshwater sediments.
Si
te
Denitri fi
(nmol -gdw
cation
Temp
Depth
(cm)

[NO -3
M
+ NO-2]
Chesapeake Bay, USA
Sta.
Sta.
Sta.
Sta.
Sta.
Sta.
Sta.
Sta.
Limf jord
Izembek
Sta.
Sta.
Sta.
1
2
5
6
7
8
9
10
, Denmark
Lagoon, USA
'1
4
10
1.
6.
4.
123.
20.
1.
25.
33.
5.0 -

0.
0.
1.
4
3
2
3
8
4
1
6
39.9

13
05
2
17
17
17
17
17
17
17
17
5

11 - 15
11 - 15
11 - 15
0 - 1
0 - 1
0 - 1
0 - 1
0 - 1
0 - 1
0 - 1
0 - 1
.5
.5
.5
.5
.5
.5
.5
.5
1.
2.
2.
29.
3.
2.
2.
5.
4
9
2
8
7
9
1
9
(OW)*
(OW)
(OW)
(OW)
(OW)
(OW)
(OW)
(OW)
Reference
This Study

















Stfrensen 1978b

0 -
0 -
0 -

2
2
2

5
2
15

(

PW)*

lizumi et

al. 1980
(PW)
<
PW)


Bering Sea
Sta.
Sta.
Sta.
12
14
19
1.
0.
1.
4
91
3
2.5
2.5
2.5
0 -
0 -
0 -
2
2
2
7.
2.
3.
9
9
9
(PW)
(PW)
(PW)
Koike and


Hattori 197!


Tokyo Bay, Japan
Sta.
Sta.
Sta.
1
2
2
2.
4.
8.
0
2
4
8.0
8.0
20.0
0 -
0 -
0 -
2
2
3
1.
5.
15
6
9

(PW)
(PW)
(Exp)
Koike and

Koike and
Hattori 197(

Hattori 197!
Mangoku-Ura, Japan
Sta.
Sta.
Sta.
Sta.

7
10
3
3

11
7.
0.
8.
29.

6
15
9
3
21
21
15
3.5
21
0 -
0 -
0 -
0 -
0 -
2
2
2
2
3
b2.
2.
b
3
34.2
22.
30.
2
0
(PW)
(PW)
(PW)
(PW)
(Exp)*
Koike and
Koike and
Koike and
Koike and
Koike and
Hattori 197!
Hattori 197!
Hattori 197!
Hattori 197'
Hattori 197!
Continued.
-------
Table 7.  Continued.
Site
Tokyo Bay, Japan
Simoda Bay, Japan
Intertidal Mudflat,
New Zealand
Louisiana Lakes, USA
Freshwater
Saline
Toolik Lake
Lake Naivasha, Kenya
Denit
(nmol


3.
0.

0
0
1.
ri fication
8.4
17.8
9 - 39.b
12 - 1.40
4.5
1.8
- 0.12
- 0.12
4 - 82.9
Temp
20
26
22
22

4
4

Depth [NCTo + NO'o]
(cm) (>l)
0 - 3 15.0 (Exp)
0 - 3 15.0 (Exp)
0 - 7.5 10UO (Exp)

0-12
0-2 13,000

Reference
Koike and Hattori 197£
Koike and Hattori 197!
Kaspar 1982
Smith and DeLaune 198!
Kl ingensmith &
Alexander 1983
Viner 1982
*OW  = Ambient concentration of overlying water
 PW  = Ambient concentration of pore water
 Exp = Experimental  concentration
-------
                                CONCLUSIONS

     These results strongly support  the  concept  that  hyperbolic  substrate-
rate relationships govern denitrification  even  for  mixed  bacterial
assemblages in natural  sediments.   However,  even  in a  single  estuarine
system for a given season,  there is  considerable  variance in  the  kinetic
parameters (Omax and Ks)  which  describe  this  hyperbolic  relation  at  a given
station.  The correlative information  presented  in  this  paper suggests  that
physical and chemical  features  of  the  sediment  and  overlying  water  are
largely responsible for the observed ranges  in  kinetic parameters for
sediment denitrification  in this estuarine system.
-------
                              LIST OF FIGURES

Figure 1.  Location of stations in the Chesapeake Bay  and  adjacent  tribu-
           taries where surface sediments were sampled to  determine
           denitrification potentials.

Figure 2.  Redox potentials (based on platinum electrode)  of  estuarine  sedi-
           ments at ten stations in Chesapeake Bay.

Figure 3.  Denitri fication potentials (based on ^0  production)  at  different
           concentrations of nitrate at selected stations  in  the Chesapeake
           Bay (Sta. 6, Ragged Pt., Sta. 7,  St. Leonards;  Sta.  9, Horn  Pt.;
           Sta. 10, Windy Hill).

Figure 4.  Denitri fication potentials (based on NgG  production)  at  different
           concentrations of nitrate at Sta. 3 in Chesapeake  Bay.

Figure 5.  A) Response of denitrification potentials at different concen-
           trations of nitrate in  estuarine  sediments  at Horn Point.  B)
           Linear transformation of data in  A) which was used to calculate
           Michaelis-Menton kinetic constants of denitrification rates.

Figure 6.  Denitrification potentials of estuarine sediments  based  on amend-
           ments of 100  vM of l^N  labelled nitrate (Twilley  and  Kemp,
           unpublished data).
-------
                 CHAPTER TWO








    SCIENTIFIC SIGNIFICANCE  AND  MANAGEMENT



   IMPLICATIONS OF SEDIMENT  OENITRIFICATION




            IN THE CHESAPEAKE  BAY
              Robert R.  Twilley



                     and



               W. Michael  Kemp
Horn Point Laboratory,  University  of  Maryland



   P. 0. Box 775,  Cambridge,  Maryland 21613
                 Final  Report



                    to  the



     U.S. Environmental  Protection  Agency



            Chesapeake  Bay Program



             Annapolis,  Maryland
-------
                                CHAPTER TWO
           SCIENTIFIC SIGNIFICANCE AND MANAGEMENT IMPLICATIONS OF
               SEDIMENT DENITRIFICATION IN THE CHESAPEAKE BAY
                                INTRODUCTION
     Various investigators have argued that the first step toward developing
a sound strategy for managing nutrient waste loading to an estuary is to
budget the major inputs and sinks of each nutrient (e.g. N, P, Si) for that
estuarine system (Nixon 1980).  Managers are particularly interested in the
fate of nitrogen in estuaries since it may be the "limiting nutrient" most
often controlling rates of algal production (Ryther and Dunstan 1971;
'Goldman et al. 1973; Boynton et al. 1982).  Accordingly, EPA developed a
preliminary budget of nitrogen for Chesapeake Bay, and although sediment
recycling processes were evaluated in this budget, nitrogen sinks or losses
in sediments were not included.  Increased loading of nitrogen to Chesapeake
Bay and its tributaries has been documented over the last several decades
(D'Elia 1982), and the question of possible sinks for elevated levels of
nitrogen is, therefore, an important management issue.

     Denitrification is an obligate anaerobic process wherein nitrate (NO;}),
which is required as an electron acceptor, is transformed to nitrogen gas
(N£ or N20).  Since N-fixation is generally insignificant in estuaries
(Nixon 1981), gaseous nitrogen is generally not available for biological
incorporation and readily diffuses from sediments to the atmosphere.  Hence,
denitrification represents a sink in the nitrogen cycle of aquatic systems.
Nitrate can diffuse from the water column to anoxic zones within the sedi-
ments and then enter the denitrification process.  Additionally, ammonium,
which is abundant in most interstitial waters, can be oxidized to nitrate
(nitrification) in the aerobic upper layers of sediments.  This nitrate may
be denitrified after diffusing along a concentration gradient to anoxic
sites either deep in the sediments or in reduced microzones in upper layers.
In the first process, nitrate available in the water column is utilized,
while in the second nitrification is required and is coupled to denitrifica-
tion (Jenkins and Kemp 1984).  Hence, there are at least two denitrification
pathways, and the magnitude of each will determine the rate of nitrogen loss
from the system.

     In this chapter we present a preliminary budget of nitrogen loss from
the Chesapeake Bay resulting from denitrification occurring in estuarine
sediments.  Since no direct measurements of denitrification exist for the
mainstem of the Bay, this budget will be based largely on results of denitri-
fication studies on two tributaries of Chesapeake Bay funded by NOAA through
University of Maryland Sea Grant College (Contracts # NA81AA-D-00040 and
-------
NA84AA-D-00014).  These studies used  15N methodologies  to  trace  both  the
direct utilization of nitrate from the overlying  water  and nitrate  reduction
coupled to nitrification in surficial  sediments  (Twilley and  Kemp,  unpub-
lished data).  This preliminary estimate of denitrification is discussed
relative to nitrogen budgets of the Bay and to nitrogen recycling processes
in estuarine sediments.  In a latter  section of  this  chapter  we  discuss the
implications of this process relative to important  issues  of  water  quality
management.


               PRELIMINARY BUDGET OF  SEDIMENT DENITRIFICATION

     Preliminary estimates of total N loss  from  the Choptank  and Patuxent
River tributaries have been calculated (Table 1)  taking into  account
significant spatial and temporal  variations in nitrification  and denitrifica-
tion rates in estuarine sediments (Twilley  and Kemp,  Unpublished data).
Denitrification rates for any one sediment  type  are dependent on nitrate
concentrations in overlying water (Fig. 1).  Thus,  ambient rates may  vary
substantially along the length of an  estuary with decreasing  nitrate  concen-
trations from upper to lower reaches  (Fig.  2A,    NO^   amendments.   The
influence of nitrate concentration in the overlying water  on  denitrification
will also result in seasonal variation in rates,  since  peak concentrations
are associated with spring freshets and minimum  concentrations occur  during
the summer (Fig. 2B,   NO^ amendments).  Nitrification  also varies  sea-
sonally with lower redox conditions during  the summer inhibiting this
strictly aerobic process (Fig. 2B, ^NH^ amendments).  These  processes may
also vary with sediment type, suggesting the importance of spatial  hetero-
geneity of their rates (Chapter One).  We have found  that  the response of
nitrification and denitrification to  these  factors  varied  significantly
between two tributaries with different nitrogen  loading characteristics
(point vs. diffuse source inputs).

     Total nitrogen loss in the Choptank River tributary from both  direct
and coupled denitrification was estimated at 1059.4 My  y'1 (note:   Mg =
million grams).  Nitrogen loading to  this system, which is heavily  influenced
by agriculture, is estimated at 1944  Mg N y  .   Therefore, denitrification
losses represent nearly 55% of the total inputs.  Similar  results were found
for the Patuxent River which had a total N  loss  via denitrification of 543.7
Mg y   or 51% of total N inputs.

     Estimates of daily nitrogen loss were  calculated for  each respective
segment of the tributary on an areal  basis  taking into  account the  seasonal
variation in total (direct plus coupled) denitrification rates  (Table 2).
Daily denitrification rates decreased with  increasing salinity.   For
instance, denitrification in the Choptank ranged  from 7.7  to  22.1 kg
N*km~^'d~* from the mouth to the headwaters, and  for  the Patuxent these
respective rates ranged from 9.86 to  13.96  kg N-km"^-d~ .   The higher rates
in the headwaters of the Choptank River are associated  with higher  nitrate
concentrations that occur in this tributary.
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Figure 2.  A) Denitrification  rates  ftased orf^N-^

           salinity regions of  the Choptank and Pat

           the mainstem Chesapeake Bay.   The solid

           nitrate concentrations along  the length
                                          paroduction  in  three

                                         ruxent River  estuaries and

                                         line  represents  ambient

                                         of the two  tributaries.
B)  Seasonal  rates  of  denitrification (1$N_N production)  in two

salinity regions  of the Choptank River estuary.   For  both A and
 intact sediment  cores were amended with either    NO^ or    NHl,

which represent direct or coupled denitrification  rates,

respectively.
                                                                             B,
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Table 2.  Estimates of nitrogen loss  from combined  denitrification  rates
          (coupled plus direct, see Table 1)  for each  segment  of Choptank
          and Patuxent River tributaries.
Segment

Lower
Middle-A
Middle-B
Upper
.Total

Lower
Middle
Upper
Total
Area
(km2)

172.0
62.4
38.2
5.8
278.4

85.4
33.2
10.6
129.2
a t
Spring and Fall* Summer'
(My N-d"1) (Mg N) Mg N'd"1

4.042
2.3U8
1.926
0.341
8.617

2.239
1.395
0.414
4.048
CHOPTANK
48b.04
276.96
231.12
40.92
1034.04
PATUXENT
268.68
167.40
49.68
485.76
RIVER
0
0
0.130
0.039
0.169
RIVER
0.258
0.100
0.029
0.387
Total
Annual Rates
T(My N)

0
0
19.50
5.85
25.35

38.70
15.00
4.35
b8.05
(Mg N)

485.04
276.96
250.62
46.77
1059.39

307.38
182.40
54.03
543.81
Mean
Dai ly Rates
(ky N'km *d

7.73
12.16
17.97
22.09
10.42

9.86
15.05
13.96
11.53
*Seasonal rates given as daily means in million grams per day  (My N'd"1)
 and million grams per seasons (My N).

^Seasonal rates based on 120 d per spring plus fall  season.

^Seasonal rates based on 150 d per summer season.
     These areal estimates of daily denitrification rates  for the Choptank
and Patuxent Rivers were applied to segments of the Chesapeake Bay to esti-
mate loss of nitrogen from this system (Table 3).   Average rate for the
headwaters of these two tributaries was 18.03 kg N*km~^'d    and average rate
for higher salinity regions was 8.80 kg N-knT^'d"1 (Table  3).  The latter
rate was applied to the lower middle segment of the Bay located in Virginia
waters.  It is unknown what rates are applicable to higher salinity regimes,
so approximations were made based on half of the rate for  the high salinity
regions of the Choptank and Patuxent tributaries (4.40 kg  N-km"^-d~M.
Using statistics on the area of each segment, total  denitrification for the
Chesapeake Bay was estimated at 68.7 Mg N-d"1 (Table 3).
-------
Table 3.  Estimates of nitrogen loss via total  denitrification  for each
          segment of Chesapeake Bay based on average areal  rates  for  the
          Choptank and Patuxent River tributaries (Tables 1  and 2).
Segment
Northern
Middle
Upper
Lower-MD
Lower-VA
Southern
Upper
Central
Lower
Total
Area*
(km2)
738.9

1395.4
987.9
1144.9

506.6
433.8
1058.7
6266.2
Rate per
Area
(kg N-km'^-d"1)
18.03

16.51
13.61
8.80

4.40
4.40
4.40

Rate per
Segment
(My N-d'1)*
13.32

23.04
13.45
10.08

2.23
1.91
4.66
68.69
*Data from EPA (1982).

^Million grams per day (Mg N d~M.


     The estimate of nitrogen loss  from the Chesapeake Bay  via  denitrifica-
tion is about 40% of the inputs of  nitrate (plus  nitrite)  to the  water
column of the tidal Bay system (Table 4).   Based  on a total  input of new
nitrogen amounting to 336.2 Mg N d   , denitrification removes about  20% of
this loading.  In correcting annual  inputs of nitrogen fluxes into the Bay
(EPA 1982) by taking into account denitrification,  annual  total  N inputs  are
decreased to 347.5 Mg N d"1.  The most striking result is  that  nitrogen loss
via sediments (68.7 Mg N d~*) is greater than benthic recycling of ammonium
(40.0 Mg N d"1) so that the net removal is 28.7 Mg  N d"1.   These  corrections
to the estimates of average nitrogen inputs to the  Bay suggest  that  denitri-
fication may be a very significant  process to the cycling  of nitrogen within
this estuarine ecosystem.
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Table 4.  Average annual  nutrient input to the water column  of the tidal
          Chesapeake Bay  system.   Data from Table 3 and EPA  (1982).*
Nutrient

TN
NOX
NH4
TKN

TN
NOX
NH4
TKN
Atmos.
Sources

50.2
18.0
11.1
32.2

50.2
18.0
11.1
32.2
Fluvial
Sources
Denitri fication
221.3
138.5
11.3
72.8
Denitri fication
221.3
138.5
11.3
72.8
Point
Sources
Omitted
64.7
21.5
33.8
42.5
Included
64.7
21.5
33.8
42.5
Benthic
Sources

40.0

40.0
40.0

-28.7
-68.7
40.0
40.0
Total

376.2
178.1
96.2
187.5

347.5
109.4
96.2
187.5
*Rats given as million grams per day (Mg N g~M.


          PROJECT RELEVANCE TO ISSUES OF WATER  QUALITY  MANAGEMENT

     While the scientific understanding of estuarine nutrient  cycling  has
improved dramatically in the last several  decades,  this development  has  been
paralleled by a general  intensification of water  quality management  problems.
Degradation of nutrient  water quality conditions  in the Chesapeake and other
estuaries calls for improved interaction between  estuarine  scientists  and
resource managers.  Some of the issues related  specifically to denitrifica-
tion include the following:

     1.  Is full tertiary treatment of sewage effluents really necessary?

     2.  What is the impact of agricultural  NO^ losses?  To what extent  do
         they need to be controlled?

     3.  Would nitrification of sewage effluents  facilitate natural  removal
         of N from the estuary by allowing the  direct shunt of N0§ into
         sediment denitrification?
-------
     4.  Do the dynamic cycles of estuarine nutrient-productivity relations
         suggest seasonally differing strategies for treatment of sewage
         effluents?

     5.  Where should such effluents be located in the estuariane gradient
         for maximum cost-effectiveness and minimum ecological impact?

     In general, an improved understanding of benthic nutrient processes
will benefit resource managers concerned with nutrient waste treatment and
with fisheries.  Excessive nutrient enrichment of estuaries can lead to
reduced water quality and degraded fisheries yields; however, the sediment
process, denitrification, can ameliorate this problem via natural nutrient
removal pathways.  The question is -- when, where and to what extent do
estuarine sediments affect nitrogen removal, and how are these processes
influenced by various environmental factors?

                                   Anoxia

     A majority of the nitrate that is microbially reduced to nitrogen gas
in sediments of the Choptank and Patuxent tributaries is produced from
nitrification occurring in these sediments.  This coupled nitrification-
denitrification process varies seasonally and spatially depending on redox
conditions of surface sediments, owing to the strict aerobic requirements
for nitrification.  In the spring cool water temperatures result  in rela-
tively high oxygen solubility and low respiration rates.  Thus, redox
potentials in surficial layers of sediment are suitable for nitrification.
During the summer, higher sediment respiration rates, together with lower
solubility of oxygen in the overlying water, result in reduced depth-
penetration of 02 into sediments, thereby limiting bacterial nitrification.
In general, it appears that when nitrification rates are high ammonium
regeneration from sediments is low, whereas when nitrification rates decline
in the summer, ammonium regeneration rates increase (Henriksen and Kemp,
1986).  Therefore, the absence and presence of sediment nitrification may  be
a key process determining the relative amount of nitrogen that is recycled
or lost from estuarine ecosystems.

     The susceptibility of nitrification to redox conditions suggests that
dissolved oxygen concentrations in water overlying sediments may  strongly
influence nitrogen fluxes in estuarine sediments.  In areas of the Bay where
anoxia exists, coupled denitrification in pelagic sediments may be inhibited
and result in an increase in ammonium flux out of the sediments.   In a study
of water quality parameters along a lateral transect across the Bay at
38°33.5' N latitude, periodic anoxic events in the shallows were  associated
with elevated NH^ concentrations and subsequent increases in phytoplankton
biomass (Malone et al. 1986).  These periodic anoxic events may occur in a
time scale of a few days and, depending on the response of sediment nitrifi-
cation to absence of 02 in the overlying water, may alter the exchange of
nitrogen across the sediment-water interface.

     Besides these periodic anoxic events in the shallows, deeper bottom
waters of the mainstem Bay are anaerobic for extended periods during the
summer.  The summertime occurrence of anoxia in bottom waters of  the
partially stratified Chesapeake Bay has been documented since the late
-------
1930's (Newcombe and Home 1938).   However,  its  temporal  and  spatial  dimen-
sions appear to have increased in  recent  years  in  response  to  cultural
eutrophication (Officer et al. 1984).   The  amount  of  water  in  the  main  part
of the Bay that is devoid of oxygen has increased  fifteen  fold from  195U  to
1980 (EPA 1983).  If increased nutrient levels,  particularly  nitrogen,  are
responsible for this poorer water  quality condition,  then  the  presence  of
anoxia may be exacerbating this problem by  indirectly increasing  nitrogen
recycling from the sediments.

     There are a variety of direct and indirect  detrimental consequences  of
these anoxic conditions on the estuarine  benthos (Kemp  and  Boynton 1981;
Officer et al. 1984; Seliger et al. 1985).   A fundamental  step in  describing
and understanding' this seasonal phenomenon  is the  development  of  dissolved
oxygen budgets.  In particular, the major processes  contributing  to  total 02
demand need to be partitioned  into habitat  (e.g.,  water column vs. benthos)
and biotic groups (e.g., microbial vs. metazoan).   Two  recent  reports con-
sidering anoxia in the Bay conflicted  in  their  analyses of  02  budgets,  where
one (Taft et al. 1980) attributed  most of the 02 demand to  water  column
processes while the other (Officer et  al.  1984) concluded  that benthic
respiration dominated.  Direct measurements  have shown  that both  benthic  and
planktonic 02 demand can be important  (Kemp  and  Boynton 1980;  Boynton and
Kemp 1985).  During spring, it has beens  suggested that nitrification may be
significant to the total sediment  oxygen  demand  (SOD) whereas  during  summer
SOD is strongly influenced by  the  oxidation  of  reduced  by-products of sedi-
ment respiration.  At present, there are  insufficient data  available  to
determine which metabolic processes are important  in  these  02  budgets;  such
information is essential toward EPA's  goal  of developing predictive  models
for water quality management.

     It appears then that processes associated  with  nitrification  and deni-
trification in estuarine sediments are certainly important  to  the  issue of
anoxia in the Chesapeake Bay.   Nitrification may contribute to oxygen con-
sumption in sediments and thus a component  of oxygen  budgets  of the  Bay.
Conversely, anoxia may inhibit nitrification, thus decreasing  nitrogen  loss
via nitrification-denitrification  coupling  resulting  in an  increase  in
ammonium recycling from estuarine  sediments.

                   Mainstem Bay vs. Estuarine Tributaries

     This study presents preliminary evidence that denitrification poten-
tials are higher in estuarine  sediments located  in the  tributaries compared
to the mainstem of the Chesapeake  Bay.  In  a study of the  Patuxent and
Choptank River tributaries using ^N methodologies and  intact  sediment
cores, a strong relationship was found between  ambient  denitrification  rates
and N03 concentrations in the overlying  water  (Fig.  1).  These rates were
used to consider the importance of dentrification  in  the nitrogen  budgets of
these two systems (Table 1).  While these rates  and  budgets probably  repre-
sent reasonable estimates for  these and related  tributaries of the Bay,
their relevance for the mainstem Bay is unclear.  Thus, the calculatons of
denitrification for the mainstem Chesapeake  Bay  (Table  3)  may  be
overestimated.
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     This suggestion that denitrification potentials of sediments in the Bay
are lower than in the tributaries  is relevant also to understanding the loss
of N via the coupling of sediment  nitrification to denitrification.  As
demonstrated for the Patuxent River estuary, more than 90% of the NO^ pro-
duced from nitrification in the sediments is denitrified during the spring
(Jenkins and Kemp 1985).  In the mainstem Bay, this coupling may be weaker
since the affinity for nitrate reduction in these sediments is lower.  This
coupled reaction is very important to total N£ production in both Patuxent
and Choptank tributaries, but its  significance in the mainstem Bay may be
lower.

                               Ambient Rates

     The study described in this report on denitrification potentials in
estuarine sediments of Chesapeake  Bay has important implications to locating
those sites in this estuary with optimum nitrogen removal capabilities.
Maximum denitrification potentials occur in the oligohaline regions of the
tributaries of the Bay and are associated with areas having high carbon
content and low N:P ratios of surficial sediments.  At higher salinities,
denitrification potentials are lower, probably resulting from lower nitrate
concentrations occurring in these  regions.  This observation coincides with
the rapid uptake of nitrate in these regions as evident from plots of
nitrate versus salinity (e.g.  Kemp and Boynton 1984; l/Jard and Twilley
1986).

     It is difficult to convert denitrification potentials measured in this
study to ambient rates (per sediment surface area) because nitrate diffusion
in sediment pore-waters must be taken into account.  However, the spatial
distribution of denitrification potentials established here provides a
rational basis for selecting sites for more extensive measurements of ambient
denitrification rates in the Bay.   Direct measurements of ambient nitrifica-
tion and denitrification rates in  mainstem Bay sediments would be required
to obtain dependable assessments of the importance of these processes in
ameliorating cultural eutrophication.  Direct measurements of ambient nitri-
fication and denitrification (such as those available for the Choptank and
Patuxent sediments) are also necessary for rigorous calibration of water
quality models.  Our preliminary N-budget for the mainstem Bay indicates 40%
of the nitrate loading and 20% of  the total nitrogen loading is lost via
denitrification.  The data for denitrification potentials provided here
(Chapter 1, Table 4) indicate that these N-removal rates may be slightly
overestimated; however, other regions of the Bay not considered (such as
Susquehanna Flats) may exhibit highly active denitrification rates, meaning
our budgeted N-losses would be underestimated.

     Obviously, the question of denitrification loss as a term in the estua-
rine N-budget is central to the developoment of an effective water quality
management program for Chesapeake  Bay.  This report provides preliminary
estimates to address that question; definitive, unequivocal answers would
require direct empirical measurements.
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                               FIGURE LEGENDS

Figure 1.  Response of ambient denitrification  rates  (based  intact  core
           measurements using li5N labelled nitrate)  to nitrate concentrations
           in the overlying waters.

Figure 2.  A) Denitri fication rates based on   N-^  paroduction in  three
           salinity regions of the Choptank and Patruxent  River estuaries  and
           the mainstem Chesapeake Bay.  The solid line represents  ambient
           nitrate concentrations along the length of the  two tributaries.
           B)  Seasonal rates of denitri fication (^N-N production) in two
           salinity regions of the Choptank River estuary.   Eor both A and B,
            intact sediment cores were amended  with  either   NO^ or   NH^,
           which represent direct or coupled denitrification rates,
           respectively.
-------
                                 REFERENCES
  1.   Andersen, J.M.   1976.  Rates of denitrification of undisturbed sediment
      from  six  lakes as a  functon of nitrate concentration, oxygen, and
      temperature.  Arch.  Hydrobiol. 80:8-21.

  2.   Balderston, W.L., B. Sherr and W.J. Payne.  1976.  Blockage by acety-
      lene  of nitrous  oxide reduction in Pseudpmgnas perfecto-marinus.
      Applied and Environmental Microbiology 31:504-508"^

  3.   Billen, G.  1975.  Nitrification in the Scheldt estuary.  Estuarine
      Coastal and Marine Science 3:79-89.

•  4.   Billen, G.  1978.  A budget of nitrogen recycling in North Sea sedi-
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-------
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-------
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-------
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-------
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APPENDICES
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