Limnol. Oceanogr., 38(2), 1993, 314-330
0 1993, by the American
Society of Limnology
and Oceanography,
Inc.
Coupling of microbial fermentation and respiration processes
in an intertidal mudflat of the Elbe estuary
M. Kerner
Institute
for Hydrobiology
and Fishery Science, Zeiseweg 9, D-2000 Hamburg
50, Germany
Abstract
In laboratory experiments, the consumption of terminal electron acceptors was studied under defined
redox conditions during different seasons. Under oxic conditions oxygen together with nitrate and Mn
was reduced. Under suboxic conditions nitrate reduction occurred simultaneously with Mn and ferric Fe
reduction and the fermentation of organic matter. Over an annual cycle, maximum reduction rates were
found in early summer. These were 24.7 1 for oxygen, 1.13 for Mn(IV) and 3.33 pmol g-l d-l for Fe(III).
In winter, the respective rates decreased 3.4-, 4.7-, and 9.2-fold. Nitrate reduction remained constant
from July to November at about 5.62 cLmo1 g- ’ d-l and decreased 2%fold by February. The production
of CO, due to fermentation of organic substances was the same magnitude as that for oxygen respiration
in summer. When fermentation processes stopped at the end of summer, nitrate respiration out-competed
Fe and Mn reduction for organic substrate. These results indicate that even great differences in free energy
do not prevent coexistence of different respiration processes in the same sediment layer. Microbial
reduction rates in the sediment appear to depend on organic matter of low molecular weight which is
produced during fermentation.
Current ideas on the microbial consumption
of terminal electron acceptors in sediments
have been greatly influenced by the conceptual
model of Billen and Vanderborght (1976), who
interpreted concentration profiles in sediments
and first postulated that the reduction processes are separated from each other in individual sediment layers. It was argued that the
electron acceptors are preferentially used according to their free energy and are thus consumed in the following sequence: O2 4 N03x Mn(IV) -+ Fe(III) -, SO,*- --t fermentation
(Froelich et al. 1979; Zehnder and Stumm
1988).
There is evidence that some of these respiration processes coexist in the same layer. This
discrepancy has been explained by microhabitats, i.e. spaces of more reduced conditions
within a spatial structure of a sediment layer
(Jorgensen 1977). That the reduction of nitrate
is independent of the number and size of anoxic microhabitats in sediments was shown by
Sexstone et al. (1985). Robertson et al. (1988)
demonstrated microbial
aerobic denitrification or corespiration of nitrate and oxygen.
Lovley and Phillips (1988) found that bacteria
isolated from oxic freshwater sediments reduced Fe(III) to Fe(II) in the presence of
Acknowledgment
This work was supported by the Physico-Chemical
partment, GKSS-Forschungszentrum,
Geesthacht.
De-
Mn(IV). They hypothesized that Mn(IV) oxidation of Fe(II) is the mechanism that ultimately prevents Fe(II) accumulation
in the
Mn(IV)
reduction
layer. Westermann
and
Ahring (1987) found that sulfate reduction,
methane production,
and denitrification
occurred simultaneously in the same soil horizon
of a swamp, and Lovley and Phillips (1987)
described the competitive mechanisms among
those processes. These findings make it questionable that free energy alone controls the coexistence or separation of microbial reduction
of different electron acceptors.
The distribution
of various electron acceptors might be explained by the fact that most
respiration processes depend on low-molecular-weight organic matter. These compounds
are formed as end products of fermentation
processes (McInerney 1988). Thus, any quantification of the fermentation by the formation
of organic matter is erroneous when microbial
fermentation is coupled with respiration. The
same is true for C02, which is an end product
of both respiration and fermentation.
According to Froelich et al. (1979), sedimentary environments
are classified as oxic
with microbial oxygen reduction being the predominant process. When oxygen is absent and
nitrate, Mn, and ferric Fe reduction occurs, the
environment is suboxic. The anoxic environment is characterized by microbial sulfate reduction. The present work was done to study
314
315
Sediment microbial interactions
6
2
0
5b
m]
plant
---Z
El
mud
Fig. 1.
Location
250
260
debris
-----z-zEl
mj
impermeable
horizon
m
0 a0 stones
larP 0
of the sampling site in the Elbe estuary and on a cross-section
the environmental conditions under which microbial respiration processes coexist. An analytical device was used that continuously exchanges the pore water in undisturbed sediment
zones of an intertidal mudflat sediment under
controlled laboratory conditions. Artificial pore
water containing nitrate was continuously percolated to model a suboxic environment during submersion. Aerobic water containing nitrate was used to simulate an oxic environment
during exposure to air. The consumption of
terminal electron acceptors was monitored
continuously in the pore water after its passage
through the undisturbed sediment structure.
Microbial
respiration
processes were measured by changes in the concentrations of 02,
N03-, NO*-, Fe(II), Mn(II), and S*-. Fermentation processes were calculated from the
Xi0
fine/coarse sand
of Heuckenlock.
difference between the concentration of CO,
measured in the pore water and the production
of CO, as computed from respiration processes.
Materials and methods
Study site- Sediment samples were collected from an intertidal, freshwater mudflat at
Heuckenlock, a wildlife reserve along the Sfiderelbe, upstream from Hamburg Harbor (Kerner et al. 1986). The region drains into the Elbe
only via a tidal channel, and consequently the
sediments are very homogeneous in their
physical characteristics. The study site, on a
bank -5 m from the tidal channel, is covered
by Phragmites communis stands (Fig. 1). The
abundance of the burrowing benthic organisms, mainly Oligochaeta (1.80% Naididae
316
Table 1. Variations
sediments.
Kerner .
in the input parameters used in the different phases of the percolation
Season
Phase
Duration
00
(mg Zer-
Feb 88
1
2
3
113
187
65
0.0
0.0
37.0
May 89
1
192
15.6
Jun 88
1
2a
2b
3
Ju189
1
145
Nov 88
1
2
3a
3b
3c
4
5
140
72
117
47
97
168
46
14.5
20.9
4.4
37.2
I)
Additive
with mudflat
NO,-N
NH,-N
(mg liter-l)
30
30
30
6.5
0.0
6.0
1.0
1.0
1.0
70
4.0
5.0
50
50
50
50
2.0
7.0
7.0
7.0
1.0
1.0
1.0
1.0
0.0
20
7.0
5.0
0.0
15.6
37.0
37.0
37.0
0.0
0.0
50
50
50
50
50
50
50
6.0
6.0
6.0
6.0
6.0
6.0
6.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
0.0
36.1
36.1
0.0
and Tubificidae),
changes over the year between 200 and 2,000 individuals dm-*.
At the sediment surface, 8 1% of the intertidal mud consists of silt and clay (163 pm).
The water content (wt/wt) of the sediment is
generally 70% and the organic matter content
is 30% of the solid material. Nitrate concentrations measured by in situ pore-water analysis never exceeded 1.02 mg N liter- l. The pH
of the water covering the sediments during high
tide varied between 7.1 and 7.8, depending on
the time of year. In the same period, nitrate
ranged between 0.3 and 5.1 mg N liter-l, dissolved oxygen was 0.3-8 mg liter- l, and the
temperature in the sediment surface (-2 cm)
was between 1.5 and 17.4”C.
The mudflat sediment at the study site is
covered by ~80 cm of water for -3 h and
exposed to air for - 9 h during each tidal cycle.
During this period, the mean rate of percolation is 2.56 x 1O-4 ml s-l cm-2 (Kerner et al.
1990).
Sampling and incubation of sediment coresDuring different seasons in 198 8 and 1989,
undisturbed sediment cores were sampled to
a depth of 10 cm with Plexiglas tubes (6.4-cm
i.d.) and a special corer (Kerner et al. 199 1).
The samples were taken while the tidal flats
were exposed at low tide. The ends of the sediment cores were sealed with parafilm to minimize evaporation and immediately placed in
jars free of oxygen (Becton Dickinson and Co.,
-
Pressure
(cm HO
experiments
ATU
ATU
NaOAc
GasPak System). Within 1 h of sampling, the
cores were transported in the absence of oxygen at ambient temperatures to the laboratory
for further treatment.
In a glove box, the sediment cores were cut
in a pure nitrogen atmosphere (0, _( 0.05% of
saturation) into 2-cm layers at precut spots in
the Plexiglas tube. The sediment surface layer,
still in its Plexiglas ring, was then placed in the
percolation unit. The percolation unit is part
of a device constructed to allow a controlled
percolation of synthetic pore water of known
chemical composition, analog to that in situ,
through the sediment disk. The construction
and function of the apparatus has been described elsewhere (Kerner et al. 199 1). A pressure equivalent to a precalculated water column was used to force the synthetic pore water
through the sediment layer (Table 1). The pressures used were always less than that exerted
by the water column in nature. Therefore, possible anomalies caused by pressure were minimized.
The parts of the percolation device in contact with the pore water were made of Plexiglas, and the feed pipes were made of Teflon
or Viton to minimize adsorption. Before the
experiments began, the whole apparatus was
soaked in 5% HCl overnight. Deionized water
(Milli-Q;
Millipore)
was always used for dilution and rinsing.
The synthetic pore water contained am-
Sediment microbial interactions
monium [(NH&SOJ,
nitrate (NaNO,) (see
Table I), and 600 mg liter- 1 sulfate (Na,SO,).
Chloride (CaCl,) was dissolved in the synthetic
pore water to a concentration of 600 mg Clliter-’
to permit better polarographic
measurements and to produce a difference in concentration in the in situ pore water of the sediments, which was always above the chlorinity
of -200 mg Cl- liter- 1 in the natural pore
water of the sediment samples. However, the
concentrations of all ions in the synthetic pore
water were within the range of natural changes
at the study site (ARGE 1990). All percolation
experiments were conducted under a pure nitrogen atmosphere at 12°C. The use of stable
temperatures and synthetic pore waters represents a compromise between simulation of
natural conditions and the control and reproducibility
of the experiments. The synthetic
water free of total inorganic C (ZCOJ had a
pH of 9.6 and 6.7 for NHQ-N concentrations
of 1 and 5 mg liter-l, respectively.
During the experiments, different phases
(Table 1) were created by varying the concentrations of 0, in the synthetic pore water by
purging with a gas mixture of N2 and 0,. An
oxygen concentration of up to 37 mg liter-’
was obtained by purging with pure oxygen. In
the last phase of one percolation experiment,
100 mg liter- 1 of sodium acetate was added to
the synthetic pore water as a substrate for respiring organisms in the sediment (Table 1).
Allylthiourea
(ATU), a metal chelator with a
high affinity to cuprous Cu was used under oxic
conditions at a concentration of 10 mg liter-l
(Table 1) to inhibit nitrification
of ammonia
to hydroxylamine
by the copper enzyme ammonia-monooxygenase
(Hooper and Terry
1973).
At the beginning of the experiments, the flow
rate was measured with a fraction sampler and
reported as the amount of water that had passed
through the sediment layer during 5-min intervals. The pore water thus obtained was analyzed for chloride. While the experiments were
being conducted, samples of at least 5 ml of
pore water were collected in the absence of
oxygen in calibrated polyurethane syringes. For
the duration of sampling, the flow rate was
recorded. The flow rates in the experiments
reached up to two times those measured in
situ. These are given in Table 2 as the mean
volumetric flow rates for the whole period of
317
each different experimental phase. The pore
volume affected by percolation
was never
~50% of the total.
The pore-water sample was transferred into
a Teflon cell for polarographic analyses, always
maintaining
suboxic conditions.
After the
electrochemical measurements had been completed, the sample was filtered through a cellulose acetate filter (Minisart, PS 0.45 pm) and
stored in a polyethylene bottle (10 ml) at - 20°C
for further analyses.
Analytical procedures for dissolved compounds-Concentrations of sulfide, ferrous Fe
and bivalent Mn were determined by differential pulse polarography with a static Hg electrode in the unfiltered pore water sampled during the percolation
experiments.
The
instrument used was a Metrohm, 647 VA polarograph and 646 VA processor (Davison
1979; Luther et al. 1985). The scan rate was
10 mV s-l with a pulse amplitude of 50 mV
and a drop time of 600 ms. This procedure
allowed H2S, HS-, and S2- to be determined
together. The detection limit for S2- was 1 pg
liter- *; for Fe2+ and Mn2+ it was 0.05 mg liter- l.
The samples used in electrochemical determinations were analyzed for Cl-, NO3 -, N02-,
NH4+, and POd3- in a flow injection analyzer
according to the methods described in the
manual (Tecator, Aquatec 5020). Sulfate was
detected by ion chromatography
(Dionex
2000i). The detection limit was 5 mg liter-l
for chloride, 1 pg N liter- 1 for NO;?-, 30 pg N
liter- 1 for N03-, 10 pg N liter- 1 for NH4+, 5
pg P liter- 1 for POd3-, and 0.1 mg liter- 1 for
so42-.
Chloride was also determined in the in situ
pore water during every percolation experiment. The sediment in the upper 2 cm was
homogenized and weighted with an inert fluid
(FC 72, 3M-Germany)
in tubes designed for
centrifugation to obtain the pore water (Batley
and Giles 1980). After centrifugation the pore
water was removed in a syringe, filtered through
a cellulose acetate filter (PS 0.45 pm), and stored
for further analysis at -20°C.
Measurements of oxygen, pH, and CO2 were
made with a glass flow-through cell containing
specific electrodes (Eschweiler and Co., Kiel).
Pore water was sucked through this cell immediately after its passage through the sediment to exclude any losses of gases by diffu-
=
p3
0.00~0.00
0.00~0.00
0.01
- 12.7kO.33
0.01
-28.9kO.20
0.01
-23.6kO.47
1.00
-22.6k1.91
0.09
0.00zk0.00
1.00
o.oo+_o.oo
o.oo+o.oo
-
Nov 88 (Vol 53, t, = 1,200)
1-suboxic
0.55-to.17
PI =P2
0.01
2-oxic
0.2220.07
0.01
P2 = P3a
0.41kO.14
3a-oxic
0.01
P3a = p3b
0.27kO.17
3b-oxic
1.00
p3b = p3c
0.24kO.13
3c-oxic
1.00
P3c = P4
0.14kO.06
4-suboxic
0.01
F4 = Ps
5 -suboxic
0.36kO.07
0.77kO.12
0.01
Jul 89 (Vol = 23, t, = 1,200)
1 -suboxic
1.45-t0.32
3 -suboxic
p2b
0.00~0.00
0.01
-8.10
-
- 15.3zkO.20
May 89 (Vol = 60, t, = 1,200)
1- oxic
2.36-1-0.24
Jun 88 (Vol = 36, t, = 800)
1-suboxic
1.55k0.16
0.01
PI = P2a
0.36-tO.08
2a-oxic
0.01
p2a = p2b
1.25kO.42
2b-oxic
1.oo
0.00~0.00
0.01
-35.3kO.24
o.oo+o.oo
Feb 88 (Vol = 58, t, = 2,700)
1 -suboxic
51kO.13
0.01
PI = P2
2 -suboxic
0.1540.10
1.00
P2 = P3
0.19+0.10
3-oxic
Phase/
significance
-5.69k0.12
0.01
-1.31kO.17
0.01
+0.19~0.10
0.01
-2.17kO.07
0.01
+0.48+_0.08
0.01
-5.58kO.26
1.00
-5.92kO.11
-2.83kO.10
- 1.65k0.12
0.01
-0.15kO.21
0.01
-0.85kO.01
0.01
- 1.74kO.03
-3.73kO.10
-6.05kO.12
0.01
o.oo-to.oo
0.01
-2.25 50.20
NO,-N
‘)
-0.28kO.17
0.01
-0.53kO.22
1.00
-0.66kO. 12
1.oo
+0.43+0.17
0.12
-0.71+0.16
0.03
-0.22kO. 11
0.01
+0.45 k0.23
+0.03-tO.20
+0.24+0.04
0.01
-0.19kO.15
1.oo
-0.10~0.05
0.01
+0.23*0.07
-0.39kO.33
+ 1.22kO.33
0.01
+0.71*0.17
0.01
-0.97 kO.3 1
(mg liter-
NH,-N
0.95&O. 10
0.01
0.04kO. 10
1.oo
o.oo_+o.oo
1.oo
o.oo_+o.oo
1.oo
o.oo_+o.oo
0.01
0.29
0.02
1.48
0.45kO.07
7.33k1.51
0.01
0.00~0.00
1.00
o.oo+o.oo
1.00
0.03 kO.02
4.22kO.47
4.43kO.59
0.03
3.68kO.22
0.01
0.06+-O. 10
0.98 kO.20
0.01
0.58kO.17
0.03
0.44 kO.08
1.00
0.44 kO.06
1.00
0.39kO.07
0.01
0.59kO.28
0.01
2.07kO.77
0.99f0.22
2.44k0.3 1
0.01
0.38
1.oo
0.32
0.01
0.54kO.06
2.67kO.39
9.92kO.60
0.01
4.77kO.10
0.01
4.OOkO.24
I)
nd
nd
nd
nd
nd
nd
2.23kO.37
1.4OkO.28
nd
nd
nd
1.8OkO.30
nd
13.5k5.80
0.01
71.3k17.1
nd
nd
(jig liter-
S’-
6.73k0.43
0.01
6.97&O. 14
0.01
6.75kO.13
1.00
6.73kO.08
1.00
6.76kO.06
1.00
6.62kO.09
0.01
7.13
6.64kO.14
6.75kO.04
0.01
7.34kO.03
0.01
8.00+0.01
0.01
8.2OkO.16
6.84-1-O. 14
7.92kO.20
0.06
7.47&O. 10
1.oo
7.46a0.40
PH
0.04
0.06
0.29
0.29
0.16
0.20
0.21
0.19
1.08
1.04
1.04
0.27
0.27
0.15
0.19
0.079
Flow
(ml min-‘)
Table 2. The mean consumption (-) and production (+) of chemicals and the physical parameters under stable conditions during suboxic and oxic
percolation through the surface layer of an intertidal mudflat sediment in different seasons. [+SD (6, _ ,); pX = pX + , 2 0.05, mean values are not
significantly different; a pore volume (Vol, ml) was exchanged each t, minutes; not detected-nd].
Sediment microbial interactions
sion. The detection limit for oxygen was 0.01
mg liter-l. The sum of the CO2 in all dissolved
ionic forms was calculated from the pH and
the ionic strength of the pore water. The detection limit for dissolved CO2 depended on
pH but was never BO.03 mM.
15Ntracer experiments- Stable nitrogen isotopes were used as a tracer during the experiments in July and May; they were added to
the synthetic pore water in the form of N03(Amersham, 98 atom% 15N) and NH4+ (98.1
atom% 15N), respectively. NH4+ was eliminated from 3-ml subsamples of the pore water
used for calorimetric analysis by purging with
N2 after raising the pH to 9.2 with a borate
buffer. The NH3 gas which emerged was
trapped at pH 2 in 3 ml of 0.04 N H,SO,. The
different fractions of the subsample containing
(R) and NH4+ (E) were anN03-/N02-/DON
alyzed for 15N and 14N atom% after transformation of the nitrogen species to N2 at 10m4
torr according to the Dumas method (Yoneyama and Kumazawa 1974) by emission spectroscopy (Jasco N 150).
Due to suboxic conditions in the experiment
of July 1989, changes in the 14N : 15N ratio
resulted only from the formation of dissolved
organic N (DON) during percolation.
319
organic matter, CM (pm01 ml- l), was calculated from the isotope dilution of dissolved NH4+
(Laws 1984):
CiM =
ln[(Ei15 - a15)/(E,15 - CX’~)]
ln(CtIC,A)
(3)
where Eil’ and ES15are the 15N atom% in the
fraction (E) of the pore-water sample and the
synthetic pore water. Cl” and CsA are the respective NH, + concentrations km01 ml- ‘).
The arithmetic mean of DON, PON, and of
the ammonium mineralized, CM, was used in
Eq. 6 for (C - s) to compute production rates
by Eq. 7.
Calculation of rates-The changes in the
concentrations of chloride measured after the
synthetic pore water passed through the sediment layer are explained by the displacement
of in situ pore water from individual pores in
the sediment by the synthetic pore water. The
mass eluted (E) during percolation was calculated with Eq. 4:
E=
i=n(Cj + Ci+*)
2
4
_
s
x
Cvi +
i=l
x
cc+
1 -
vi+l)
2
6)
(4
Ri14 _ (CiDoN Ck!14)
+ (CiN@14)
(1) where Cj (pg ml- ‘) is the concentration of chloRi15 (CiNpl’) + (CiDoNa15)
ride measured in the outflow for the increment
of measurement i = 1, 2, 3 . . . n at time ti
where Ri14and Ri15are the 14N and 15N atom%
(min) and the volumetric
flow rate Vi (ml
in the fraction (R) of the sample with an inminl);
S
(pg
ml-l)
is
the
concentration
of the
crement of measurement i = 1,2,3 . . . n; cy14y15
and @14,15
are the natural and amended levels
of 14N and 15N atom%, respectively. CiDoN
(pm01 ml-l) is the concentration of DON in
the pore-water sample, and CiN is its N03-/
NOz- concentration (pmol ml-l). The solution of Eq. 1 gives
C-DON
=
I
GN[(Ri14P15)- (Ri15P14)1. c2J
(Ri15a14)- (Ri14,15)
An equivalent of DOC was calculated from
the DON produced during suboxic percolation
in July 1989 using the C : N ratio of 15 determined for dissolved organic matter in interstitial waters of organic-rich sediments (Krom
and Sholkovitz 1977).
In the percolation experiment of May 1989,
according to the Blackburn-Caperon
model,
the production of NH4+ during degradation of
chemical in the synthetic pore water.
The exchange of chloride was calculated from
data obtained at the beginning of every percolation experiment with Eq. 4 having assumed the initial conditions:
t* = 0;
c, = co;
V’= v2,
and the final condition:
ICj - Sl 5 S/100;
i=n
where Co (hg ml-l) is the concentration
of
chloride measured in the in situ pore water of
the sediment layer. The time interval for one
complete exchange of the pore water in the
layers and its replacement by synthetic pore
water is defined by t,.
The volume of the pores in the sediment,
Vol (ml), that were affected during percolation
was computed with Eq. 5:
320
Kerner
(5)
Stable conditions for the production or consumption of a chemical in the sediment during
percolation are reached when changes in the
concentrations measured in the outflow are
small. For the calculation of the onset of stable
conditions with constant concentrations,
the
data recorded during the whole experiment
were fitted to a straight line on a graph (Press
et al. 1986) after eliminating the first value on
the time scale and repeating the process until
a desired significance value was reached. With
a value of x 2 5 0.2 and a minimum of five
data points, the regression analysis was significant at the P < 0.01 level.
The arithmetic mean c and the standard
deviation 6,- 1 of the concentrations were calculated for the different chemicals under stationary conditions. Student’s t-test was used
to statistically compare the arithmetic means,
Px = P,+~, for the same chemicals, calculated
for successive phases of a single percolation
experiment (Press et al. 1986). At P L 0.05,
the mean values were not statistically different.
The production or consumption of a chemical during passage of the synthetic pore water
through a sediment layer is expressed by Eq. 6:
A mineralization
rate of NH4+ was calculated
with the measured X02 production assuming
the Redfield ratio of C : N = 6.7 for the organic
material completely degraded during the percolation experiments.
CO, production was equated to the microbial reduction of terminal electron acceptors
by Froelich et al. (1979), and the following
molecular relationships were found: 1.25 for
1 for 02, 0.5 for Mn4+, 0.25 for Fe3+,
NW,
and 2 for S042-.
From the reduction rates of the different processes measured during percolation, a production rate (pm01 g- 1 d- ‘) for ZCO, can be calculated with Eq. 9, assuming
complete
degradation of the organic matter to CO,:
PC= z: Is” x f”
m-i
where f m is the equivalent of X0,
produced
per mol of the electron acceptor m and Zsmthe
reduction rate given in Eq. 8 for the i-different
acceptors.
A degradation index was calculated by Eq.
10 to determine the efficiency of degradation
of organic matter to X0,
in the microbial
respiration processes:
Q=(&S)o
(6)
where c (pm01 ml-‘) is the mean concentration of the chemical in the pore water measured at the outflow and calculated as described above, ij (ml min-‘)
the mean
volumetric flow rate during percolation, and
S (pm01 ml-‘) the concentration of the chemical at the inflow.
The reactions take place within the pore volume affected by percolation and calculated in
Eq. 5. The rate (pm01 ml- 1 d- ‘) of the sum of
the processes that lead to the concentration of
the chemical in the pore water is given by Eq. 7:
(7)
The water content divided by the dry weight
was calculated as being Q,= 2.33 for the intertidal mudflat sediment. A reaction rate (pm01
g-l d-l) bas ed on the dry weight of the sediment is given by Eq. 8:
I, = Ip x 4.
(8)
where Z,” is the production rate for CO2 computed by Eq. 8 with the data for X0,
measured during percolation. When the factorsf”
given above are used to calculate P, with Eq.
9, their dependency on differing composition
of the organic material is disregarded. If the
factors for degradation of carbohydrates, lipids, and proteins are corrected to those described by Anderson et al. (1986) then the degradation indices theoretically
would range
between 1.0 and 1.33.
The computer program MINEQL (Westall
et al. 1976) was used to calculate whether compounds in the pore water measured during different phases of percolation were oversaturated or undersaturated
with respect to the
minerals. When undersaturated, a dissolution
from the solid could not be ruled out. The pH
values in Table 1 were used at the ionic strength
resulting from the composition of the synthetic
pore water for the calculations discussed here.
Calculations were made for the cations Na+,
Sediment microbial interactions
K+, Ca2+, Mg2+, Fe2+, and Mn2+ and the inorganic ligands C032-, HC03-, P043-, OH-,
Cl-, S042-, and HS-. Stability constants for
FeSO,, FeOH+, and FeCl+ which were different from those used in the program were taken
from Davison (1979).
Results
In general, in the initial phase 1 of the percolation experiments, synthetic water free of
oxygen and amended with N03- was used.
Oxic conditions began in phase 2 (June, November) or phase 3 (February). Thereafter,
suboxic conditions were re-established at phase
3 (June) and phase 4 (November) (see also Ta-
ble 1).
pH, which ranged from 6.6 to 8.2 during
percolation experiments (Table 2) did not differ much from the pH (6.8-7.8) of in situ pore
water of the same sediment layers (Figs. 2-4).
Thus, errors in the rate measurements due to
pH could be ruled out.
X0,--In
all experiments oxygen was first
detected in the pore water - 1 h after the beginning of oxic percolation (Figs. 2-4). After
suboxic percolation began, oxygen immediately decreased in the pore water of the sediment (Figs. 3,4). Thus, additional oxygen consumption processes must have taken place after
oxic percolation began.
The concentrations in the in situ pore water
measured at the beginning of the experiments
in phase 1 were 3-5 times those reached under
stationary conditions of phase 1. Calculations
suggested that the carbonate concentrations
measured during the different phases of percolation were always in equilibrium
with respect to CaCO,. Hence, dissolution of CaCO,
from its solid phase could be ruled out under
steady state conditions. Production rates for
CO, (Table 3) were calculated from the X0,
concentrations determined under stable conditions of the different phases (Table 2).
The degradation indices, D1, computed with
Eq. 10 were never > 1.3 under oxic conditions.
If we assume no precipitation
of CaC03, this
value is reached only when all organic substance has been converted to C02. Likewise,
degradation indices much higher than 1, detected in the previous phase under suboxic
conditions, clearly indicate that additional CO,
must have been produced during processes dif-
321
ferent from the respiration of inorganic electron acceptors.
Indices as low as 0.14, which were also calculated for oxic conditions, can be explained
only by an incomplete oxidation of organic
matter. That such low indices were not due to
precipitation
of CaCO, was revealed during
the subsequent phase when suboxic conditions
were re-established in the sediment layer (Table 2). During that phase, DI increased to a
value of 0.66 although pH decreased.
Nitrogerz-The
dissolved inorganic forms of
nitrogen are given in Figs. 2-4 as differences
between concentrations in the pore water before and after its percolation through the sediment layers. The horizontal dashed lines mark
the level at which both concentrations were
the same. Concentrations below this line indicate gross consumption of the compound and
concentrations above gross production.
Constant N.03- concentrations were reached
in all experiments (Figs. 2-4), and remained
unchanged for more than a week of continuous
suboxic percolation. Thus, under these experimental conditions limitation of NO,- reduction in the sediment layers by organic material
could be ruled out for all seasons. The N03reduction rates under suboxic conditions, both
at the beginning and the end of percolation,
indicate that N03- reduction during percolation was limited by NO,- if concentrations in
the outflow were eO.4 mg N liter-l (Tables 2,
3: June phases 1 and 3; November phases 1
and 4).
After the transition from suboxic to oxic and
back to suboxic conditions, NO,- concentrations changed without an obvious time lag and
constant concentrations
were reached after
about one complete exchange of pore water
(Figs. 2-4). When concentrations
of oxygen
were between 0.3 and 2.9 mg liter- l, N03concentrations
were reached that indicated
gross consumption of this compound (Table
2). Gross production of N03- was observed
only at high oxygen concentrations
(> 8 mg
liter- l) when nitrification
exceeded N03- reduction (Figs. 2-4).
When nitrification
was inhibited by ATU,
the NO3 - concentrations
steadily decreased
and stable concentrations were reached after
about one complete exchange of pore water.
This inhibition was reversible, and N03- con-
322
Kerner
,
phase : 1
0”
z,
7
I
I”
z,
7
i
I
-3.1
NO3- N : 0
NH4- N : A
Non- N : q
I
I
I
-4.8
-6.5
I
0
I
38
I
I
76
r‘
t
I
c,
I
I
114
I
152
I
I
I
190
I
228
I
I
266
I
I
304
I
Ill
342
380
time ( h )
Fig. 2. Concentrations of different chemicals measured in the pore water after passage through the surface layer of
a mudflat sediment in February. Conditions in the inflow during phase 1 -anaerobic
with N03-; phase 2-anaerobic;
phase 3-aerobic
with N03-.
Sediment microbial interactions
8.5
phase : 1 ,
I
2a
323
3
,2b,
I I
6.5
Fe (II)
Mn (II)
I=?
:0
:o
=6
5
E-3
TI
5
.-
fO
E
z
IO
.
6
0.2
z,
z
I
-0.6
I"
z
-1.4
4
I
a,
i
Non-N:
0
2
'c-9-2.2
0
7
<3 -3.0
time ( h )
Fig. 3. As Fig. 2, but in June. Conditions in the inflow during phase 1 -anaerobic
with NO,-, phase 2b-with
ATU; phase 3-anaerobic
with NO,-.
with NO,-;
phase 2a-aerobic
324
Kerner
8
I
Q7
6
13.24
0” -0.4
z,
z -1.8
‘m-4.6
0
3
-6
0
70
140
210
280
350
420
490
560
630
700
time ( h )
Fig. 4. As Fig. 2, but in November. Conditions in the inflow during phase 1 -anaerobic
aerobic (15.6 mg 0, liter-l) with NO,-; phase 3a, c-aerobic
(37 mg 0, liter-l) with N03-;
phase 4-anaerobic
with NO,-; phase 5-anaerobic
with NO,- and acetate.
with NOa-; phase 2phase 3b-with
ATU;
325
Sediment microbial interactions
Table 3. The degradation indices D, calculated from reduction rates (pm01 g-’ d-‘) of the electron acceptors and
CO, production under oxic and suboxic conditions in the surface layer of a mudflat sediment in different seasons.
Season/
phase
02
NO,-
Fe(II)
Mn(H)
.S
Feb 88
1 -suboxic
2 -suboxic
3-oxic
0.00
0.00
9.59
1.969
0.00
1.391
0.36
0.72
0.09
0.82
0.95
0.63
0.0019
0.0244
0.00
May 89
1 - oxic
7.21
4.026
1.14
0.74
0.00
1.13
0.68
0.57
0.99
x0*
D,
2.32
1.65
1.65
0.78
2.34
0.14
35.68
2.76
0.0014
0.00
0.00
0.00
39.40
35.25
122.68
78.28
7.66
1.33
4.81
Jun 88
1 -suboxic
2a- oxic
2b-oxic
3 -suboxic
0.00
24.71
0.00
2.98
1.04
5.93
12.61
3.33
0.00
0.00
0.06
Jul89
1 -suboxic
0.00
5.61
0.22
0.50
0.0012
40.35
5.51
0.00
5.05
9.15
13.56
12.96
0.00
0.00
5.42
1.19
+0.14
2.86
0.63
1.52
1.07
0.23
0.09
0.00
0.00
0.00
0.02
0.07
0.24
0.13
0.08
0.15
0.13
0.04
0.10
0.0008
0.00
0.00
0.00
0.00
0.00
0.00
7.34
2.80
4.17
4.98
4.42
0.53
0.92
1.05
0.42
0.45
0.29
0.32
0.28
0.66
Nov 88
1 -suboxic
2 - oxic
3a-oxic
3b-oxic
3c-oxic
4 - suboxic
5 -suboxic
centrations increased immediately
after removal of ATU from the synthetic pore water
to a level similar to that before inhibition (Figs.
3 and 4).
During different seasons with oxygen concentrations of between 1.7 and 28 mg liter-l,
N03- reduction was 47-7 1% of that found
during suboxic conditions (Table 3) as calculated from N03- concentrations during inhibition of nitrification.
Oxic incubation periods
of up to 333 h did not inhibit the capacity of
the sediments for N03- reduction.
Throughout
the experiments,
N02- remained undetectable when stable N03- concentrations were observed (Figs. 2-4). Hence,
during this period N02- and N03- reduction
occurred at the same rate.
Under suboxic conditions, when N03- reduction increased > 25-65%, NOz- accumulated when N03- concentrations decreased. Up
to 67% of the N03- additionally reduced was
then detected in the form of NOz- at concentrations which never were > 1.2 mg N liter- I,
indicating that NOz- reduction followed zeroorder Michaelis-Menten
kinetics in the range
between 0 and 1.2 mg N liter-l. After transition from suboxic to oxic conditions, NOzaccumulated to concentrations always < 0.7 mg
.
N liter-’ during the period of increasing N03concentrations. During this period, oxidation
of NH4+ to N02- must have been preferred
to the step of oxidation of NOz- to N03-.
Rates measured with ‘jN--At the beginning
of percolation (phase l), NH4+ concentrations
in the outflow became constant within the time
required for one complete exchange of the pore
water (Figs. 2-4). However, NH4+ concentrations in the synthetic pore water used in percolation experiments were always different
from those detected in the in situ pore water
measured at the beginning of phase 1. I therefore conclude that adsorption-desorption
processes of NH 4+ between pore water and the
solid phases in the sediment were either completed during one exchange of the pore water
or remained undetectable due to continuous
flow of the pore water. Likewise, changes between suboxic and oxic conditions were immediately followed by changes in NH,+ concentrations
(Figs. 2-4). This observation
provides further evidence that the NHdmbconcentrations reached during different phases of
percolation
depended mainly on microbial
processes that were induced under these specific conditions. Hence, from the stable NH4+
concentrations and the abundance of 15N, re-
326
Kerner
Table 4. Production of DON and NH,+
percolation experiments.
calculated
from the abundance of the 15N isotope at different
Jul 89
Time
R’S
W
(atom%)
20
45
116
142
90.7
81.8
86.4
88.7
times of the
May 89
C”“N
(mg liter-‘)
0.39
0.76
0.57
0.45
@ON = 0.54kO.16
liable values for the mineralization
of NH,+
were calculated (Table 4). These rates were 75%
of those calculated from X0, production (Table 2).
Metals-Equilibrium
calculations suggested
that Fe(II) concentrations measured in the pore
water during percolation were always undersaturated with respect to minerals. Thus, precipitation
of Fe(II) could probably be discounted. The same was true for Mn reduction.
During percolation with anaerobic water
containing N03- (phase l), stable concentrations of Fe(II) and Mn(I1) were obtained after
the time required for one complete exchange
of the pore water (Figs. 2-4). Under suboxic
conditions, when N03- was removed from the
sediment layer (Fig. 2) rates of Fe reduction
increased by a factor of about two while Mn
reduction rates remained unaffected (Table 3).
In the presence of dissolved oxygen, Fe(II) concentrations rapidly decreased below the detection limit. However, Fe reduction was recorded even when oxygen was present
at
concentrations of -0.3 mg liter- l (Table 2).
Mn reduction decreased only between 56 and
77% when oxic conditions were introduced.
Upon re-establishment of suboxic conditions
in the sediments, it took 36-70 h of continuous
percolation of anaerobic water amended with
N03- before Fe(II) was detected in the pore
water (Figs. 3 and 4). Thereafter, concentrations steadily increased but steady state conditions were not reached even after 175 h of
continuous percolation. Constant concentrations of Mn were detected within 70 h after
suboxic conditions began (Figs. 3 and 4). The
acceleration of Fe and Mn reduction shortly
after addition of acetate (Fig. 4) showed that
both processes had been limited by organic
substrates in the previous phase and not by
the availability
of Fe(III) and Mn(IV).
Time
(h)
21
45
69
164
190
E’S
(atom%)
15.3
39.0
7.5
17.0
38.3
CM
(mg liter-l)
8.27
3.95
9.73
7.68
3.91
@ = 6.71f2.64
SOd2- reduction as evidenced by the appearance of sulfide in the pore water was not
accompanied by changes in SOd2- concentrations. Sulfide at no stage was detected in the
pore water when suboxic conditions were reestablished. The differences in the rates at the
beginning and at the end of percolation were
therefore not produced by precipitation of FeS
and MnS.
Discussion
Coexistence of respiration processes-The
results demonstrate that under suboxic conditions the reduction of N03-, Fe(III), Mn(IV),
and SOd2- coexisted in the two-dimensional
spatial structure of the sediment surface of the
freshwater mudflat sediment from the Elbe estuary. When conditions were changed from
suboxic to oxic in this sediment, oxygen reduction was immediately induced at high rates,
while Fe reduction became completely inhibited. This finding could have been due either
to the inhibitory effect of oxygen on Fe reduction or to the chemical oxidation of Fe(II)
(Ahonen and Tuovinen 1989). N03- respiration was inhibited by only 47-7 1%, even under
conditions of oxygen saturation. Similarly, Mn
reduction decreased by 56 and 77% at high
concentrations of oxygen. Mn(I1) oxidation was
too slow to affect the rates measured during
percolation (Yeats and Strain 1990).
These findings indicate that organic matter
degradation might be severely underestimated
when different terminal electron acceptors are
assumed to be consumed in individual
sediment layers. Furthermore, the coexistence of
respiration processes is important in understanding the cycling of trace metals in sediments (Kerner and Wallmann 1992).
It must be concluded from the high rates of
NO3 - and Mn reduction under oxic conditions
327
Sediment microbial interactions
that these processes were not entirely restricted
to suboxic conditions. Aerobic N03- and Mn
reduction may be advantageous to microorganisms that must adapt to fluctuating oxicsuboxic conditions in some sediments (Kuenen and Robertson 1988; Nielsen et al. 1990;
Lloyd et al. 1987). In the sediment surface of
intertidal-flat
sediments the changes in redox
conditions might occur even within a single
tidal cycle (Kerner et al. 1990). In this type of
sediment the capacity to reduce oxygen could
be interpreted as a mechanism for decontamination in order to prevent any inhibition
of
anaerobic respiration processes. This conclusion is supported by the percolation experiments which revealed that specific respiration
processes are induced immediately after transition from suboxic to oxic conditions.
Rate measurements-During the annual cycle, rates of oxygen reduction were highest in
early summer and N03- reduction rates remained high from July to November in the
sediment surface of the intertidal mudflat sediment as observed with the percolation technique (Fig. 5). Similar seasonal variations have
been described by Christensen et al. (1990) in
the sediments of a nutrient-rich
Danish lowland stream. In past suspension experiments
of subtidal sediments from different regions of
the Elbe estuary, changes in NO,- reduction
did not correlate with the seasons (Wolter et
al. 1985). The rates detected varied between
1.12 and 22.5 pmol g-l d-l and were thus
within the same range as those measured by
the percolation technique.
As reviewed by Lovley (1991), it has been
shown that most of the Fe(II) and Mn(I1) resulting from Fe(III) and Mn(IV) reduction in
aquatic sediments is not manifested as concentration changes in the dissolved phase. Fe
reduction rates as measured by Fe(II) accumulation in anoxic marine sediment slurries
have been shown to be up to 3.12 pmol g-l
d-l (Sorensen 1982). This rate is in good agreement with the Fe(II) reduction rate determined
in the summer percolation experiments. Wallmann (199 1) calculated rates of the same order
of magnitude: 18.2 pmol g- l d- l in suspensions of the same type of sediment as that used
in the experiments described here. He carried
out the experiment in summer under suboxic
conditions by measuring Fe(II) concentrations
in both dissolved and solid forms. From this
62.5
I-l,
,
m
organics
I
02
kiis.7 NOsx IO-'
I
Fe(III)x 10 ’
txxs Mn(IV)x IO-’
Feb
May
Jun
Jul
Nov
Fig. 5. Seasonal variation of the microbial reduction
rates of different terminal electron acceptors in the surface
layer of a mudflat sediment from the Elbe estuary.
comparison, the error due to formation of solid
forms of Fe(II) in poorly defined phases, which
were not included in the equilibrium
calculations, could result in a 5.5-fold underestimation. Although this underestimate seems quite
large, Fe(III) reduction still remains low with
respect to CO2 production.
The calculation of SOd2- reduction by production of dissolved sulfides did not take
chemical interactions in the sediment into account, and the rates given here are minimum
values. However, SOd2- was never significantly consumed during the percolation experiments. Thus, SOa2- reduction may be excluded as a major process resulting in the
formation of CO2 under the conditions of the
experiments.
Signijicance of fermentation -As described
above, the respiration rates given here can be
used to calculate CO, production with Eq. 9
without major errors. Furthermore, clear evidence was obtained that all the X02 in the
pore water was produced during microbial
degradation of organic material. Thus, excess
X02 production can be explained only by fermentation processes and was used as such to
quantify this process.
If we assume dissolution of carbonate from
the solid phase when CO, was produced, the
following equation applies:
CaCO, + CO2 + H20 + Ca2+ 4 2HC03-.
Thus, in the worst case, fermentation rates calculated for the percolation experiments would
be overestimated by a factor of 2. Likewise,
only degradation indices >2 calculated from
Eq. 10 are indicative of fermentation process-
328
Kerner
Winter
Summer
1
particulate
organic
matter
organic
matter
a
Hydrolysis
Iron and Manganese
Iron and Manganese
re;gtio*
Sulfate respiration
co, (0.02 %)
(11 %)
Fig. 6. Pathways of the degradation OForganic matter and the production
sediment from the Elbe estuary during winter and summer.
es. However, fermentation
might have been
underestimated due to microbial CO, reduction (Anderson et al. 1986).
Instead of fermentation, if the excess X0,
had been produced during Fe(III) or SOd2- reduction (which was probably underestimated
as discussed above), the specific rates of these
reductions would have been about 600-fold
higher than those measured. These rates would
be much higher than measured previously
(Lovley 199 1).
Data on DON provides further support for
my estimates of fermentation. The production
of DOC during percolation calculated from
DON was found to balance about half the CO2
produced by fermentation.
In the field, humification and adsorption might act as a sink
for fermentation products (Rashid 198 5). The
finding from my experiments that dissolved
organics are not always maintained at low levels by anaerobic respiration processes is supported by previous studies of marine and limnetic pore-water profiles (Orem and Gaudette
1984; Barcelona 1980; Molongoski and Klug
1980).
The consumption of the different electron
acceptors is given in Fig. 6 in terms of biotic
\cr-\
CO, (10 %)
Sulfate respiration
co* (0.01 %)
of CO, in the surface layer of a mudflat
CO, production during both oxic and suboxic
conditions. In winter, when fermentation
is
negligible, organic material is mineralized
mainly by oxygen and N03- reduction. In
summer, CO, production by fermentation exceeds the anaerobic respiration processes by
-2-fold
and equals oxygen reduction. Similarly, an equivalent amount of organic substances of low organic weight, which could be
used as a substrate in respiration processes, is
produced.
Until now, the observation that in situ CO,
release exceeded oxygen uptake of a subtidal
sediment up to 4-fold (Hargrave and Phillips
198 1) was explained only with anaerobic respiration processes. The difficulty with this explanation is that it requires the availability
of
labile organic compounds. The percolation experiments suggest that fermentation
coupled
to anaerobic processes, which has not been
described in many sedimentary studies, might
have been overlooked.
Degradation of organic matter-Hines
et al.
(199 1) and Lovley and Phillips (1986) demonstrated that Fe(III) reduction had the potential to be a major pathway in the organic
matter decomposition
in anoxic sediments
Sediment microbial interactions
from freshwater, brackish-water, and salt-water sites. However, as shown here, in freshwater, intertidal-flat
sediments, even if the Fe
and Mn reduction rates were high they never
accounted for > 11% of the total degradation
of organic matter. Unlike the sediments described previously, the intertidal-flat
sediment
used here has a high input of oxygen and N03due to pore-water movement via percolation,
and microbial reactions are not limited by inorganic electron acceptors (Kerner et al. 1990).
This finding might explain the divergence of
my results from those obtained for sediments
where loading with electron acceptors is restricted to diffusion processes.
Fe-reducing bacteria are capable of outcompeting Sod-reducing bacteria for organic
substrates (Lovley and Phillips 1987; Sorensen
and Jorgensen 1987). In addition, the results
of my experiments also revealed that Fe reduction dominated SOd2- reduction in the sediment surface layer year-round. The inhibitory
effect of N03- on sulfide production (Jenneman et al. 1986) was not separable from its
limitation by organic substrate.
Reduction rates of Fe(III) and Mn(IV) increased during an annual cycle when substrate
became available through fermentation
processes. The maximum rate for microbial Fe
reduction found in June was followed by a
large decrease that could not be explained by
the competition
of SOd2- or Mn reduction.
N03- reduction, which reached its highest rates
when Fe and Mn reduction rates decreased,
suggests that at this time of year N03- reduction out-competed Fe and Mn reduction for
the same organic substrate. At the end of the
year, when fermentation processes were negligible, Fe and Mn reduction remained low.
Likewise, oxygen and N03- reduction processes, which can also use organic material of
high molecular weight (Nedwell 1984), were
maintained at elevated rates. In the field, where
N03- concentrations are not always as high as
in my experiments, metal reduction might be
more important during a larger part of the year.
References
AHONEN,L., AND 0. H. TUOVINEN. 1989. Microbiological oxidation
Appl. Environ.
ANDERSON,L. G.,
tion measured
Oceanogr. 31:
of ferrous
Microbial.
iron at low temperatures.
55: 3 12-3 16.
AND OTHERS. 1986. Benthic respiraby total carbonate production. Limnol.
3 19-329.
329
ARGE.
1990. Arbeitsgemeinschaft
fur die Reinhaltung
der Elbe, Wasserglitedaten der Elbe von Schnackenburg bis zur See- Zahlentafel. ISSN 093 l-2153.
BARCELONA,M. J. 1980. Dissolved organic carbon and
volatile fatty acids in marine sediment pore waters.
Geochim. Cosmochim. Acta 44: 1977-1984.
BATLEY, G. E., AND M. S. GILES. 1980. A solvent displacement technique for the separation of sediment
interstitial waters. Contam. Sediments 2: 10 l-l 17.
BILLEN,G., ANDJ. P. VANDERBORGHT.1976. Evaluation
of exchange fluxes of materials between sediments and
overlying waters from direct measurements of bacterial activity and mathematical analysis of vertical
concentration
profiles in interstitial waters, p. 154165. In Biogeochemistry of estuarine sediments. Proc.
UNESCOSCOR
Workshop.
CHRISTENSEN,
P. B., L. P. NIELSEN,J. SBRENSEN,AND N.
P. REVSBECH. 1990. Denitrification
in nitrate-rich
streams: Diurnal and seasonal variation related to
benthic oxygen metabolism. Limnol. Oceanogr. 35:
640-6 5 1.
DAVISON,W. 1979. Soluble inorganic ferrous complexes
in natural waters. Geochim. Cosmochim. Acta 43:
1693-1699.
FROELICH,P. N., AND OTHERS. 1979. Early oxidation of
organic matter in pelagic sediments of the eastern
equatorial Atlantic:
Suboxic diagenesis. Geochim.
Cosmochim. Acta 43: 1075-1090.
HARGRAVE,B. T., AND G. A. PHILLIPS. 198 1. Annual in
situ carbon dioxide and oxygen flux across a subtidal
marine sediment. Estuarine Coastal Shelf Sci. 12: 725737.
HINES, M. E., D. A. BAZ~LINSIU,J. B. TUGEL, AND W. B.
LYONS. 199 1. Anaerobic microbial biogeochemistry
in sediments from two basins in the Gulf of Maine:
Evidence for iron and manganese reduction. Estuarine
Coastal Shelf Sci. 32: 3 13-324.
HOOPER,A. B., AND K. R. TERRY. 1973. Specific inhibitors of ammonia oxidation in Nitrosomonas. J. Bacteriol. 115: 480-485.
JENNEMAN,G. E., M. J. MCINERNEY,AND R. M. KNAPP.
1986. Effect ofnitrate on biogenic sulfide production.
Appl. Environ. Microbial. 51: 1205-12 11.
JBRGENSEN,
B. B. 1977. Bacterial sulfate reduction within reduced microniches of oxidized marine sediments.
Mar. Biol. 41: 7-17.
KERNER,M., H. KAUSCH,AND M. KERSTEN. 1986. Effect
of tidal action on the distribution
of nutrients and
heavy metals in flat sediments of the Elbe estuary.
Arch. Hydrobiol.
Suppl. 75 (I), p. 118-131.
-,
AND H.-D. KNAUTH. 199 1. A new method to study biogeochemical
processes in sediments
by a percolation technique. Estuarine Coastal Shelf
Sci. 32: 173-186.
AND G. MIEHLICH. 1990. The effect of
tidal action on the transformations
of nitrogen in
freshwater tidal flat sediments. Arch. Hydrobiol. Suppl.
75 (2), p. 25 l-27 1.
-,
AND K. WALLMANN. 1992. Remobilization events
involving Cd and Zn from intertidal flat sediments in
the Elbe estuary during the tidal cycle. Estuarine
Coastal Shelf Sci. 35: 37 l-394.
KROM, M. D., AND E. R. SHOLKOVITZ. 1977. Nature and
reactions of dissolved organic matter in the interstitial
330
Kerner
waters of marine sediments. Geochim. Cosmochim.
Acta 41: 1565-1573.
KUENEN,J. G., AND L. A. ROBERTSON.1988. Ecology of
nitrification
and denitrification,
p. 161-218. In J. A.
Cole and S. J. Ferguson [eds.], The nitrogen and sulfur
cycles. Cambridge.
LAWS,E. 1984. Isotope dilution models and the mystery
ofthe vanishing 15N. Limnol. Oceanogr. 29: 379-386.
LLOYD, D., L. BODDY, AND K. J. P. DAVIES. 1987. Persistence of bacterial denitrification
capacity under aerobic conditions: The rule rather than the exception,
FEMS (Fed. Eur. Microbial.
Sot.) Microb. Ecol. 45:
185-190.
L~VLEY, D. R. 1991. Dissimilatory
Fe(III) and Mn(IV)
reduction. Microbial. Rev. 55: 259-287.
-,
AND E. J. P. PHILLIPS. 1986. Organic matter
mineralization
with reduction of ferric iron in anaerobic sediments. Appl. Environ. Microbial.
51: 683689.
1987. Competitive mechanisms for
-,
AND-.
inhibition of sulfate reduction and methane production in the zone of ferric iron reduction in sediments.
Appl. Environ. Microbial. 53: 2636-264 1.
1988. Manganese inhibition of mi-,
AND -.
crobial iron reduction in anaerobic sediments. Geomicrobial. J. 6: 145-155.
LUTHER, G. W., III, A. E. GIBLIN, AND R. VARSOLONA.
1985. Polarographic analysis of sulfur species in marine porewaters. Limnol. Oceanogr. 30: 727-736.
MCINERNEY,M. J. 1988. Anaerobic hydrolysis and fermentation of fats and proteins, p. 373-4 15. In A. J.
B. Zehnder ted.], Biology of anaerobic microorganisms. Wiley.
MOLONGOSKI,J. J., AND M. J. KLUG. 1980. Anaerobic
metabolism of particulate organic matter in the sediments of a hypereutrophic
lake. Freshwater Biol. 10:
507-518.
NEDWELL,D. B. 1984. The input and mineralization
of
organic carbon in anaerobic aquatic sediments. Adv.
Microb. Ecol. 27(7): 93-l 3 1.
NIELSEN,L. P., P. B. CHRISTENSEN,
N. P. REVSBECH,AND
J. SBRENSEN.1990. Denitrification
and oxygen respiration in biofilms studied with a microsensor for
nitrous oxide and oxygen. Microb. Ecol. 19: 63-72.
OREM,W. H., AND H. E. GAUDETTE. 1984. Organic matter in anoxic marine pore water: Oxidation effects.
Org. Geochem. 5: 175-181.
PRESS,W. H., B. P. FLANNERY,S. A. TEUKOLSKY,AND W.
T. VETTERLING. 1986. Numerical recipes: The art
of scientific computing. Cambridge.
RASHID, M. A. 1985. Geochemistry of marine humic
compounds. Springer.
ROBERTSON,
L. A., E. W. J. VAN NIEL, R. A. M. TORREMANS,AND J. G. KUENEN. 1988. Simultaneous nitrification
and denitrification
in aerobic chemostat
cultures of Thiosphaera pantotropha. Appl. Environ.
Microbial. 54: 28 12-28 18.
SEXSTONE,A. J., N. P. REVSBECH,T. B. PARKIN, AND J.
N. TIEDJE. 1985. Direct measurement of oxygen
profiles and denitrification
rates in soil aggregates. Soil
Sci. Sot. Am. J. 49: 645-65 1.
SBRENSEN,
J. 1982. Reduction of ferric iron in anaerobic,
marine sediment and interaction with reduction of
nitrate and sulfate. Appl. Environ. Microbial. 43: 3 19324.
AND B. B. JBRGENSEN.1987. Early diagenesis in
sediments from Danish coastal waters: Microbial activity and Mn-Fe-S geochemistry.
Geochim. Cosmochim. Acta 51: 1583-l 590.
WALLMANN,K. 199 1. Die Fiihdiagenese und ihr Einflul3
aufdie Mobili&
der Spurenelemente As, Cd, Co, Cu,
Ni, Pb und Zn in Sediment- und Schwebstoff-Suspensionen. Ph.D. thesis, Univ. Hamburg. 195 p.
WESTALL,J. C., J. L. ZACHARY,AND F. M. M. MOREL.
1976. MINEQL,
a computer program for the calculation of chemical equilibrium
composition
of
aquateous systems. Mass. Inst. Technol. Dep. Civ.
Eng. Tech. Note 18.
WESTERMANN,P., AND B. K. AHRING. 1987. Dynamics
of methane production, sulfate reduction, and denitrification in a permanently waterlogged alder swamp.
Appl. Environ. Microbial.
53: 2554-2559.
WOLTER, K., H.-D. KNAUTH, H.-H. KOCK, AND F.
SCHROEDER.19 8 5. Nitrification
and nitrate reduction in water and sediment of River Elbe. Vom Wasser
65: 63-80.
YEATS, P. A., AND P. M. STRAIN. 1990. The oxidation
of manganese in seawater: Rate constants based on
field data. Estuarine Coastal Shelf Sci. 31: 1 l-24.
YONEYAMA,T., AND K. KUMAZAWA. 1974. Simple determination of nitrogen- 15 in plant material by emission spectrometry. J. Soil. Sci. Plant Manure 45: 480482.
ZEHNDER,A. J. B., AND W. STUMM. 1988. Geochemistry
and biogeochemistry
of anaerobic habitats, p. l-38.
In A. J. B. Zehnder ted.], Biology of anaerobic microorganisms. Wiley.
Submitted: 20 September 1991
Accepted: 21 May 1992
Revised: 27 October 1992