Flow extremes and benthic organic matter shape themetabolism of a headwater Mediterranean stream
VICENC ACUNA,* ADONIS GIORGI , † ISABEL MUNOZ,* URS UEHLINGER ‡
AND SERGI SABATER §
*Department of Ecology, Faculty of Biology, University of Barcelona, Barcelona, Spain†CONICET, Departamento Ciencias Basicas, Universidad Nacional de Lujan, Lujan, Argentina‡Department of Limnology, EAWAG, Switzerland§Institute of Aquatic Ecology and Department of Environmental Sciences, University of Girona, Campus Montilivi, 17071 Girona,
Spain
SUMMARY
1. Single-station diel oxygen curves were used to monitor the oxygen metabolism of an
intermittent, forested third-order stream (Fuirosos) in the Mediterranean area, over a
period of 22 months. Ecosystem respiration (ER) and gross primary production (GPP)
were estimated and related to organic matter inputs and photosynthetically active
radiation (PAR) in order to understand the effect of the riparian forest on stream
metabolism.
2. Annual ER was 1690 g O2 m)2 year)1 and annual GPP was 275 g O2 m)2 year)1.
Fuirosos was therefore a heterotrophic stream, with P : R ratios averaging 0.16.
3. GPP rates were relatively low, ranging from 0.05 to 1.9 g O2 m)2 day)1. The maximum
values of GPP occurred during a few weeks in spring, and ended when the riparian
canopy was fully closed. The phenology of the riparian vegetation was an important
determinant of light availability, and consequently, of GPP.
4. On a daily scale, light and temperature were the most important factors governing the
shape of photosynthesis–irradiance (P–I) curves. Several patterns could be generalised in
the P–I relationships. Hysteresis-type curves were characteristic of late autumn and winter.
Light saturation responses (that occurred at irradiances higher than 90 lE m)2 s)1) were
characteristic of early spring. Linear responses occurred during late spring, summer and
early autumn when there was no evidence of light saturation.
5. Rates of ER were high when compared with analogous streams, ranging from 0.4 to 32 g
O2 m)2 day)1. ER was highest in autumn 2001, when organic matter accumulations on the
streambed were extremely high. By contrast, the higher discharge in autumn 2002
prevented these accumulations and caused lower ER. The Mediterranean climate, and in
its effect the hydrological regime, were mainly responsible for the temporal variation in
benthic organic matter, and consequently of ER.
Keywords: ecosystem metabolism, flow regime, Mediterranean streams, organic matter dynamics,riparian forest
Introduction
Aquatic ecosystems and their adjacent terrestrial
environments are closely connected through a variety
of material inputs and energy flows (Hill, Mulholland
& Marzolf, 2001). These connections are particularly
Correspondence: Vicenc Acuna, Department of Ecology, Faculty
of Biology, University of Barcelona, Avgda. Diagonal 645,
E-08028 Barcelona, Spain.
E-mail: [email protected]
Freshwater Biology (2004) 49, 960–971
960 � 2004 Blackwell Publishing Ltd
tight in forested headwater streams, where riparian
forest provides large amounts of particulate organic
matter to benthic food webs enhancing respiration
processes (Cummins et al., 1989), and the shade cast
by riparian plants is an important constraint on lotic
primary production (Rosenfeld & Roff, 1991; Hill,
Ryon & Schilling, 1995). Therefore, metabolism of
headwater streams has been generally considered to
be heterotrophic (Vannote et al., 1980; Cummins et al.,
1989; Rosenfeld & Roff, 1991).
Against this general background, variability of
metabolism in forested streams has been attributed
to light availability and to the dynamics of allochtho-
nous organic matter (e.g. Minshall, 1978; Hill et al.,
1995). During the annual cycle changes in the amount
of light reaching stream beds in temperate deciduous
forests can be large and rapid, because of the annual
cycles of leaf emergence and abscission (Hill et al.,
2001). Organic matter of riparian origin may influence
stream metabolism both in terms of the contribution
of allochtonous matter as well as the influence on the
accumulation processes, which depend respectively
on riparian forest phenology and the retentive capa-
city of the stream (e.g. Smock, Metzler & Gladden,
1989; Gregory et al., 1991).
Indeed, it is true that water flow is the driving force
for organic matter transport and accumulation in
streams. Flow may have a strong impact on the
metabolism of medium sized rivers (Young & Huryn,
1996; Uehlinger, 2000), while in low order streams the
relevance of flow for the dynamics of benthic organic
matter (BOM) accumulation may be affected by local
climate, topography and geomorphology. This may be
particularly true in Mediterranean streams, where
flow extremes are common and often unpredictable
(Gasith & Resh, 1999). However our understanding of
how these temporal patterns may affect metabolism of
streams is still limited.
In this paper, we aim to determine how the seasonal
variations of water flow, light and temperature could
influence the metabolism of an undisturbed, forested,
oligotrophic Mediterranean stream. The large varia-
bility of climate conditions characteristic of Mediter-
ranean streams may influence the hydrological regime
and the input and retention of organic matter (Sabater
et al., 2001). Thus, climate variability may result in
extremely large accumulations of organic matter
during low flow or, conversely, in losses during
floods. It is our hypothesis that these differences
would be reflected in stream metabolism, and that
water flow could be a driving force for metabolism,
mainly through its effect on organic matter accumu-
lation or transport. In this study, stream metabolism
was estimated using open system measurements with
diel O2 variations being recorded using a single-
station method, over a period of 18 months.
Methods
Study site
Fuirosos is an intermittent third order stream drain-
ing a catchment area of 16.2 km2 in the Montnegre-
Corredor Natural Park, a forested range close to the
Mediterranean sea (50 km north of Barcelona, north-
east Spain). The climate is typically Mediterranean,
with mild winters and warm springs and summers.
Monthly air temperatures range from 4 �C in
December to 28 �C in July and August; during
winter, air temperatures below 0 �C are infrequent.
Precipitation mostly occurs in autumn and spring
with occasional storms in summer (Bernal, Butturini
& Sabater, 2002). Mean annual discharge was about
32 L s)1 in 2001 and 87 L s)1 in 2002, with monthly
averages ranging from 0.07 L s)1 in September in
2001 to 626 L s)1 in May 2002. Permanent flow
usually ceases from July/August to September/
October (Sabater et al., 2001). At a discharge of
30 L s)1, the stream was 3–4 m wide with depths
ranging from 0.1 to 0.2 m in riffles and 0.4 m in
pools. A 50 m long reach was selected in a riffle
section of uniform channel morphology and channel
slope (0.9%). The stream has a well developed
riparian area 10–20 m wide, which forms a closed
canopy from May to October dominated by alder
(Alnus glutinosa, L.), hazelnut (Corylus avellana, L.)
and plane (Platanus acerifolia, Aiton-Willd.). The
study reach was unreplicated, so that only tentative
conclusions about riparian influences on ecosystem
metabolism may be drawn. However, these conclu-
sions are based on data from two highly different
hydrological periods.
Physico-chemical characteristics
The water level was continuously monitored from
December 2000 to December 2002 using a pressure
transducer PDCR 1830 (Druck, Leicester) connected to
Riparian influences on stream metabolism 961
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an automatic sampler (Sigma 900 Max). Discharge
was measured on several occasions using the slug-
injection method with sodium chloride as tracer
(Gordon, McMahon & Finlayson, 1992). These mea-
surements and the corresponding stage provided an
empirical stage-discharge relationship that was used
to calculated discharge based on the water level
record (Bernal et al., 2002).
Global radiation (GLR) was continuously recorded
from June 1998 to November 2002, using a SP-1110
Pyranometer Sensor (Campbell Scientific, U.K.) con-
nected to a data logger (Campbell Scientific CR-10,
U.K.) placed outside the riparian forest, about 50 m
from the study reach. GLR was converted to PAR
according to McCree (1971). Light interception by the
canopy was measured with a LAI-2000 Plant Canopy
Analyzer (LI-COR Inc., Lincoln, NA, U.S.A.). Derived
results were converted to PAR reaching the streambed
using the empirical coefficient of refraction by the
water surface of 0.49 m)1 (averaged value of 15 sets of
measurements throughout the studied period). To
validate these calculations we measured PAR reaching
the streambed by means of an underwater quantum
sensor (sensor LI-192SA fitted to LI-250 Quantum
Meter, both LI-COR Inc., Lincoln, NA, U.S.A.). The
correlation between measured and calculated values
was highly significant (n ¼ 116, r2 ¼ 0.85, P < 0.01).
Temperature and dissolved oxygen were continu-
ously recorded (see below). The pH was measured
in situ (WTW MultiLine P4, Welheim, Germany).
Stream water for nutrient analyses was filtered in situ
through glass fiber filters (Whatman GF/F filters) and
stored at 4 �C until analysis. N-NO3 and N-NH4 were
analysed using a Skalar (Breda, The Netherlands) auto
analyser, while P-PO4 was analysed with a Perkin-
Elmer spectrophotometer (APHA, 1992).
Algal and organic matter
Samples from sand and rocks were collected every
20 days for the determination of benthic chlorophyll a
(chl a). Sandy substrata were sampled using a corer
(diameter 2 cm), collecting five cores at random in the
reach. Chlorophyll on rocks was estimated from
artificial substrata (unglazed tiles 1 · 1 cm, 40 days
colonisation), which were used as surrogates for
natural communities. The tiles were glued with
silicone to natural cobbles, and were distributed in
the study reach. On each sampling date, nine tiles were
collected at random. Both cores and tiles were trans-
ported to the laboratory and chlorophyll was extracted
with 90% acetone. Routine measurements of chloro-
phyll followed Jeffrey & Humphrey (1975). Results
were expressed as chl a m)2 of streambed by using the
percentage cover of rock and sand (see below).
Direct input of particulate organic matter was
monitored using traps suspended 0.6 m above the
stream surface. The traps consisted of a square wooden
frame (1 · 1 m) and a nylon net (mesh size ¼ 1 mm)
that formed the bottom of the trap. Lateral transport to
the channel was determined by means of 10 traps
equally distributed near the channel margins. Lateral
traps consisted of a wooden frame holding a nylon net
(mesh size ¼ 1 mm). Total input was considered to be
the sum of direct and lateral input. BOM was collected
monthly with triplicate core samples with a diameter
of 20 cm diameter (mesh size 1 mm). Organic matter
samples were dried at 105 �C to a constant weight and
combusted at 450 �C for about 4 h to estimate ash free
dry mass (AFDM). Water samples for dissolved
organic carbon (DOC, <0.7 lm) analysis were analysed
using a high-temperature catalytic oxidation Shim-
adzu TOC 5000 analyzer (Shimadzu, Rydalmere,
Australia). Carbon content was calculated by applying
the empirical factor of 2.4 to the ratio gram
AFDM : gram C (Margalef, 1983).
The macroscopic distribution of substrata within
the study reach was mapped along eleven equidistant
transects using an underwater viewer 0.4 · 0.4 m.
Within this area the following were estimated:
(i) percentage streambed covered by the different
inorganic substrata (rock, cobble, sand, gravel),
(ii) percentage streambed covered by algal patches
and (iii) percentage streambed covered by the differ-
ent types of BOM (leaves, branches or fine detritus).
Finally, the average value of all transects was also
calculated for every sampling date. These averages
were used to scale up chl a and BOM to a reach scale.
Ecosystem metabolism
Assessment of metabolism rates was based on the
single-station diel O2 method (Odum, 1956). Dis-
solved oxygen concentration was continuously recor-
ded 40 times, for about 36 h. Oxygen was measured
with an oxygen meter (WTW Oxi 340-A, Weilheim,
Germany) and temperature with a 107 Temperature
Probe (Campbell Scientific, UK). The oxygen meter
962 V. Acuna et al.
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and temperature probe were connected with a data-
logger (Campbell-Scientific CR 10-X, Shepshed,
U.K.), and data recorded at 0.5 and 15 min intervals,
respectively. The oxygen probe was air-calibrated
before each measuring interval. Both probes were
placed in the thalweg area of the stream, about 5 cm
below the water surface. The saturation concentra-
tion of O2 was calculated using temperature and
elevation above sea level, according to Buhrer (1975).
Net Production rate of O2 [b(t)] in g O2 m)2 h)1 was
calculated using:
bðtÞ ¼ DO2
Dt� KO2
ðO2sat � O2Þz ð1Þ
where DO2/Dt is the change of O2 concentration
between two subsequent measurements, z is the mean
depth (m), O2sat is the saturation concentration of O2
(mg O2 L)1) and Ks is the reaeration coefficient of O2.
From the 40 oxygen series we excluded nine because
of supersaturation during the night, probe malfunc-
tioning, or sudden flow increase.
The reaeration coefficient Ks was determined with
propane using methods reported in Genereux &
Hemond (1992). Propane and a solution of sodium
chloride (conservative tracer) were continuously
injected about 10 m upstream of the upper end of
the study reach. Conductivity was continuously
measured at 10 and 40 m downstream of the injection
point in order to determine when the steady state in
conductivity was reached along the entire reach
(Stream Solute Workshop, 1990). These conductivity
measurements were later used to estimate the lateral
inflow through the reach. Samples for volatile tracer
concentrations were collected at 0, 5, 10 and 20 and
40 m downstream of the injection point after a steady
state was achieved. This was carried out in order to
assure that the propane diffusion into the water body
was correctly completed. The propane exchange
coefficient, Kpropane was calculated as:
Kpropane ¼1
sln
G1C2
G2C1
� �ð2Þ
where s is travel time of water (min)1) between an
upstream station, at 10 m from the injection point (1)
and a downstream station, at 40 m from the injection
point (2), Gi is the steady state concentration of
propane at the respective sites, and Ci is the steady
state conservative tracer concentration at the respect-
ive sites, corrected for the background concentration.
The reaeration rate of propane was converted to
oxygen using a factor of 1.39 (Rathbun et al., 1978).
The DO changes integrate the response of a stream
ecosystem over a reach length £3v/Ks, where v is the
average velocity and Ks is the reaeration coefficient
(Uehlinger, Konig & Reichert, 2000). In Fuirosos, the
average stream length considered by the metabolism
estimates was therefore 346 m, and ranged from 50 to
1000 m. The derivative (DO2/Dt) at the time ti (day)
needed for the calculation of metabolism rates (eqn 1)
was the analytical derivative of a 2nd second order
polynomial fitted to the data. For this fit the data were
weighted with a normal distribution centred at ti and
with a standard deviation of 0.05 day. The calculation
of the derivative with the polynomial fitting technique
strongly reduces the effect of errors in O2 measure-
ments, which is greatly amplified by the differenti-
ation process (Shoup, 1983). Such errors are ‘averaged
out’ when daily rates are calculated but they strongly
affect the determination of hourly of half-hourly rates
needed to evaluate photosynthesis–irradiance (P–I)
curves. Three metabolic parameters were based on net
production rates of O2 (eqn 1): ecosystem respiration
(ER), gross primary production (GPP), and net
ecosystem production. ER was calculated as the sum
of net O2 production rate [b(t)] during the dark period
and respiration values during the light period, these
being calculated as the linear interpolation between
the net O2 production rate values of sunrise and
sunset of the nights before and after the day of
interest. GPP was the sum of net O2 production rate
during the light period and respiration rates during
the light period, as explained above. Annual rates of
GPP and ER for 2002 were obtained by numerical
integration that was based on 26 measurements.
Stepwise regression analysis was used to explore the
potential influence of season and riparian forest GPP
and ER. Factors used were discharge (Q), water
temperature (T), daily accumulated PAR (I), per cent
of cobbles cover (%), chl a (C) and BOM (OM). These
factors were not or only moderately inter-correlated.
The regression model was forced to return non-
negative output because negative values for ER and
GPP are not reasonable. The full regression model was:
X¼maxð0;aþb �Qþc�Tþd�Iþe �%þf �Cþg�OMÞð3Þ
where the dependent variable X is ER or GPP and a, b,
c, d, e, f and g are model parameters. Sub-models can
Riparian influences on stream metabolism 963
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be derived from the full model by ignoring the
influence of some factors by setting the corresponding
model parameters to zero. A backward analysis was
performed starting with the full model (3) and
iteratively eliminating the least important factor at
each step. For each dependent variable 2p¢)1 models
were tested (p¢ is the number of parameters of the full
model, light was not considered as a factor influen-
cing ER). This procedure ranked the factors for the
modelled variables GPP and ER according to their
influence. The decision, which factors should be
considered to be significant, based on the Schwarz
Bayesian Criterion (Schwarz, 1978). A detailed des-
cription of the applied regression procedure is given
by Uehlinger et al. (2000).
Photosynthesis–irradiance relationships
To evaluate the relationship between GPP and PAR
reaching the streambed, we identified the parameters
of a hyperbolic tangent function (eqn 4; Jassby & Platt,
1976) by non-linear regression (STATISTICA, version
5.5; StatSoft Inc., Tulsa, OK, U.S.A.).
GPP ¼ Pmax tanhaI
Pmaxð4Þ
where Pmax is light saturated photosynthesis, a is the
initial slope of the P–I curve (or Pmax/Ik), and I is PAR
reaching the streambed. The half-saturation light
intensity (Ik) was calculated as Pmax/a (Henley, 1993).
Results
Physicochemical characteristics
Annual rainfall was only 512 mm in 2001 (Fig. 1a). In
spite of this low input, a major flood event in January
2001 completely altered the substrata distribution in
the streambed. Later on, discharge declined almost
continuously to the end of June, when the stream
dried out (except for a few pools scattered in the
stream reach that persisted for another 2 weeks).
Flow resumed in mid October 2001. The year 2002
was wetter (annual rainfall of 850 mm). Floods
occurred between January and May, and between
September and December, and flow persisted
throughout the year. The conservative tracer injec-
tions did not indicate the accrual of water through the
study reach.
The reaeration coefficient ranged from 3.8 day)1 in
July 2002 to 187 day)1 in May 2002, and averaged
57 day)1 (Table 1). Reaeration coefficients were
strongly correlated with the product of average water
velocity and slope (r2 ¼ 0.79, P < 0.05), consistent
with the empirical relationship proposed by Tsivog-
lou & Neal (1976) to predict the reaeration coefficient.
The seasonal patterns of PAR reaching the
streambed (Fig. 1b) reflected the development of the
riparian canopy and interception by the hill slopes.
PAR increased from March to the end of April but
declined rapidly with the emergence of leaves. How-
ever, autumnal leave abscission only had a minor
effect on PAR because the hill slope caused effective
shading because of low sun angle during autumn.
Thus, except for a short period in spring, PAR was
usually <20 lE m)2s)1. Water temperatures varied
between 4 and 36 �C (Fig. 1b). Soluble reactive phos-
phorus in the stream water varied from <1 to 17 lg
P-PO4 L)1, ammonia from <1 to 45 lg N-NH4 L)1 and
nitrate from 4 to 2714 lg N-NO3 L)1. Concentrations
peaked in autumn (Table 1).
01 04 07 10 01 04 07 10 01
PA
R (
µE m
–2s–1
)
0
10
20
30
40
50
60
70
80
Wat
er t
emp
erat
ure
(°C
)
0
5
10
15
20
25
30
35
40
PARWater temperature
Dis
char
ge
(L s
–1)
0
200
400
600
800
10006000
9000
12 000(a)
(b)
Dry
per
iod
20022001
Fig. 1 (a) Discharge in Fuirosos Stream from January 2001 to
January 2003. (b) Average daily (from sunrise to sunset) pho-
tosynthetically active radiation (PAR) reaching the streambed
(continuous line) and water temperature (dotted line) in Fuiro-
sos from January 2001 to January 2003.
964 V. Acuna et al.
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Algal biomass and organic matter
Algal biomass (measured as chl a) was higher in
spring and summer and lower in autumn and winter
(ANOVAANOVA, P < 0.05). In spring 2001, chl a reached
71 mg m)2, declined after leaf emergence, and peaked
again but only in isolated pools (51 mg m)2) just
before the entire reach dried out. In the following
year, the vernal chl a peak was only 32 mg m)2,
presumably because of the preceding floods, and the
response to canopy closure was less distinct. Chloro-
phyll concentration on sand was significantly lower
than on tiles (ANOVAANOVA, P < 0.05), the values ranging
between 0.2 and 14 mg m)2.
The BOM ranged from 0.2 to 30 g C m)2, with the
lowest values occurring after the flood in January
2001 (Fig. 2). Total input (TI) peaked in autumn, at
around 2 g C m)2 day)1. Two different patterns in the
relationship between TI and BOM could be distin-
guished. In 2001, because of low water availability
during spring and summer, the litter fall started in
July and finished in November. In 2002, however,
most of the litter fall input occurred in November.
The lack of flow during summer as well as the low
discharge in autumn 2001 caused leaf accumulation
on the streambed to reach 30 g C m)2. Small debris
dams were formed, which increased water depth and
reduced current velocity. In 2002, permanent flow
and a few floods in late summer and autumn kept
BOM at low levels (maximum BOM of 10 g C m)2 in
November).
Daily and annual metabolism rates
There was significant seasonal variation in ER
(ANOVAANOVA, P < 0.05; Fig. 3). After the dry period of
summer 2001, high rates of ER (32 g O2 m)2 day)1)
coincided with large leaf litter accumulations in the
wetted channel. On the contrary, relatively low ER
rates (3.3–5.6 g O2 m)2 day)1) in autumn 2002 corres-
ponded with moderate BOM concentrations. GPP
peaked at 1.9 g O2 m)2 day)1 in spring 2002, but
subsequently declined to 0.34 g O2 m)2 day)1 when
leaves started to emerge. Seasonal variation in GPP
was less distinct, but significant (ANOVAANOVA, P < 0.05).
During winter, GPP varied between 0.1 to 1.1 g
Table 1 Physical, chemical and biological characteristics of Fuirosos, averaged on a seasonal basis. Data were obtained from January
2001 to December 2002; data from summer are those of 2002
Season Autumn Winter Spring Summer
Wetted width (cm) 215 ± 29 255 ± 2.6 241 ± 2.4 210 ± 44
Depth (cm) 11.7 ± 5.8 18 ± 3.2 23.1 ± 14.8 10 ± 11
Water velocity (m s)1) 0.13 ± 0.15 0.18 ± 0.14 0.13 ± 0.05 0.05 ± 0.01
Chlorophyll (mg chl a m)2) 7.7 ± 7.4 3.2 ± 2 15.7 ± 5.6 1.4 ± 0.8
Benthic organic matter (g C m)2) 54 ± 15 29 ± 43 23.2 ± 0.9 6.9 ± 2
Total input of organic matter (mg C m)2 day)1) 1117 ± 581 305 ± 83 506 ± 47 612 ± 102
O2 reaeration rate at 20 �C (day)1) 56 ± 26 33 ± 26 188 ± 149 5 ± 0.6
pH 7.2 ± 0.18 7.9 ± 0.13 7.6 ± 0.08 7.2 ± 0.11
Conductivity (mS cm)1) 316 ± 29.3 229 ± 24 173 ± 11.9 285 ± 51
Dissolved organic carbon (mg DOC L)1) 4.4 ± 1.7 6.8 ± 2 3.1 ± 0.54 9.8 ± 6.2
SRP concentration (mg P-PO4 L)1) 9.3 ± 10.9 3.3 ± 3.8 0.02 ± 0.007 5.7 ± 8.5
Nitrate concentration (mg N-NO3 L)1) 590 ± 161 511 ± 85 29.2 ± 0.62 3.9 ± 0.2
Ammonia concentration (mg N-NH4 L)1) 29.4 ± 4.8 15 ± 17 0.63 ± 0.07 3.4 ± 5.2
05 07 09 11 01 03 05 07 09 11 01
ER
(g
O2
m–1
day
–1)
0
5
10
15
20
25
30
35
BO
M (
g C
m–2
)
0
5
10
15
20
25
30
35ERBOM
2001 2002
Dry
per
iod
Fig. 2 Ecosystem respiration (ER, in g O2 m)2 day)1) in relation
to benthic organic matter (BOM, in g C m)2) in Fuirosos. The
non-flow period is represented by dotted vertical lines.
Riparian influences on stream metabolism 965
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O2 m)2 day)1. GPP and ER were not significantly
correlated (r2 ¼ 0.08, P > 0.05).
The annual rates of ER and GPP in 2002 were
respectively 1690 and 275 g O2 m)2 year)1, and
annual net ecosystem production equalled )1415 g
O2 m)2 year)1. P : R ratios ranged from 0.01 to 4.2
(this latter being recorded immediately before the leaf
emergence), and averaged 0.16 in 2002.
Ecosystem respiration was significantly correlated
with BOM (r2 ¼ 0.48, P < 0.05), while GPP was
better predicted by discharge (r2 ¼ 0.28, P < 0.01).
The selected model of the stepwise regression
analysis for ER, explained 60% of the variance
and included (in order of importance) BOM,
discharge and chl a as the most influential para-
meters. A three parameter model explaining 57% of
the variation was selected for GPP, with (in order of
importance) discharge, per cent of cobbles coverage
and chl a. We found no significant relationship
between daily metabolism rates and nutrient con-
centrations.
Photosynthesis-Irradiance relationships
The relationship between the instantaneous GPP rate
and PAR varied with season (Fig. 4). The relationship
was linear during summer, early autumn and late
spring (Fig. 4a). Light saturation of GPP occurred in
early spring, when PAR values were higher than
90 lE m)2s)1 (Fig. 4b). In late autumn and winter the
P–I relationship followed a temperature hysteresis
curve (Fig. 4c). The two parameters Ik and Pmax were
maximal in spring (Table 2) but differences between
seasons were not significant (ANOVAANOVA, P > 0.05),
except for Pmax which was significantly smaller
during summer (ANOVAANOVA, P ¼ 0.05). The model used
to identify these parameters (eqn 4) only provides
information on an average P–I curve because hyster-
esis is not considered.
05 07 09 11 01 03 05 07 09 11 01
g O
2 m–2
day
–1
–35
–30
–25
–20
–15
–10
–5
0
5
GPPER
2001 2002
Dry
per
iod
Fig. 3 Gross Primary production (GPP) and ecosystem respir-
ation (ER) in Fuirosos stream from June 2001 to December 2002.
0.00
0.03
0.06
0.09
0.12
0.15
0.18
0.21
PAR (µE m–2 s–1)
0
0 15 30 45 60 75 90 105
2 4 6 8 10 12 14 16 180.000
0.005
0.010
0.015
0.020
0.025
0.030
0.035
Morning valuesAfternoon values
Gro
ss p
rim
ary
pro
du
ctio
n (
g O
2 m–2
h–1
)
(a)
(b)
(c)
0 5 10 15 20 25 30 35 40 45 500.00
0.04
0.08
0.12
0.16
0.20
0.24
Fig. 4 Gross Primary production versus PAR in three typical
situations (see text): (a) 02/07/02 (summer) (b) 28/04/02
(spring) and (c) 13/02/02 (winter). Measurements in the morn-
ing are indicated as black circles and measurements in the
afternoon as white circles.
966 V. Acuna et al.
� 2004 Blackwell Publishing Ltd, Freshwater Biology, 49, 960–971
Discussion
The results obtained in Fuirosos support the view
that in oligotrophic and heavily shaded streams the
organic matter input from riparian forests is a major
determinant of ER (Webster, Wallace & Benfield,
1995). However, it is obvious in this type of system
that the seasonal effects of the changing canopy on
GPP can be extraordinary (Guasch & Sabater, 1995;
Hill & Dimick, 2002). Moreover, the hydrological
regime may significantly modify the riparian influ-
ence on ER through its effect on the amount of
allochthonous organic matter stored in the channel
(Gregory et al., 1991). This effect may be a major
driver in Mediterranean streams (Gasith & Resh,
1999). However, to reduce the limitation of an
unreplicated study similar investigations in other
Mediterranean systems are needed.
Gross primary production
In comparison with other systems, primary production
of Furiosos was relatively small (Table 3). The riparian
forest had several effects on the primary producers
within the Fuirosos. On one hand, riparian shade
conjointly with shade by hill slope at low sun angle
(from November to January), produced a clear annual
pattern of light availability (Fig. 1b). This pattern in
light availability caused relatively low GPP values for
most of the year (Table 3), except in early spring (Fig. 3)
when shade was minimal. On the other hand, detritus
covered the algal community for extended periods,
further impeding light reaching them. High discharges
had a ‘cleaning’ function by removing BOM and
allowing light to reach the primary producers, until
then partially covered by BOM. A large proportion of
the algal community in Fuirosos is made up of
encrusting red algae and cyanobacteria, which may
receive more light after the litter is removed. It is
therefore not surprising that discharge was the best
single predictor of GPP, as well as the first factor in
order of importance in the stepwise regression analysis.
Discharge has been related to GPP in other studies,
although the reasons proposed have been diverse.
Lamberti & Steinman (1997) attributed the positive
Table 2 Pmax (light saturated photosynthesis) and Ik (half sat-
uration light intensity) estimated in Fuirosos
Autumn Winter Spring Summer
Pmax (g O2 m)2
day)1)
2.8 ± 2.9 2.8 ± 2.9 3.3 ± 1.1 0.7 ± 0.3
Ik(lE m)2 s)1) 13.5 ± 9.3 13.5 ± 8.3 23.1 ± 13.1 15.3 ± 10.6
Table 3 Range or average and standard deviation of GPP, CR and P : R from low to medium size stream systems reported in the
literature
Author Method Stream
GPP
(g O2 m)2 day)1)
ER
(g O2 m)2 day)1) P : R
This study Single-station Fuirosos Stream 0.05–1.9 0.4–32.1 0.16
Bott et al., 1985 Chambers Inter-biome 0.2–3.4 0.4–2.9
Chessman, 1985 Single-station La Trobe River 0.15–1.90 3.0–4.6
Edwards & Meyer, 1987 Single-station Ogeechee River 0.5–14.0 3.7–11.5 0.25
Edwards & Owens, 1962 Single-station Ivel River 9.6 8.5 1.13
Fellows, Valett & Dahm, 2001 Two-station Rio Calaveras 0.5 0.19
Fellows et al., 2001 Two-station Gallina Creek 1.7 0.1
Fisher & Carpenter, 1976 Single-station Fort River 1.8 3.7
Hall, 1972 Single-station New Hope Creek 0.8 1.3 0.61
Hoskin, 1959 Single-station Neuse River 0.3–9.8 1.7–21.5
Kaenel, Buehrer & Uehlinger, 2000 Single-station Muhlibach 12.5 ± 4.5 8.9 ± 6.84 1.07 ± 0.08
Marzolf, Mulholland
& Steinman, 1998
Two-station Walker Branch 1.4 ± 1.8 6.5 ± 1.9
Molla et al., 1996 Single-station Montesina Stream 2.4 ± 2.1 2.2 ± 1.6 0.59 ± 0.69
Mulholland et al., 2001 Two-station Inter-biome <0.1–15 2.4–11
Uehlinger & Naegeli, 1998 Single-station River Necker 2.5 3.5 0.73
Uehlinger, Naegeli & Fisher, 2002 Single-station Hassayampa River 0.3 ± 0.1 1.65 ± 0.13 0.17 ± 0.05
Young & Huryn, 1996 Single-station Taieri River <0.3–9.6 0.7–9.8
Young & Huryn, 1999 Two-station Sutton 0.8 ± 0.3 4.6 ± 1.1 0.2 ± 0.1
Young & Huryn, 1999 Two-station Three O’Clock 3.7 ± 0.9 2.7 ± 0.9 1.5 ± 0.2
Riparian influences on stream metabolism 967
� 2004 Blackwell Publishing Ltd, Freshwater Biology, 49, 960–971
relationship between GPP and discharge to high
nutrient loading associated with high discharges.
However, Uehlinger & Naegeli (1998) and Uehlinger
(2000) reported that high discharges could shift
ecosystem metabolism towards heterotrophy because
benthic primary producers are more susceptible to the
abrasion by moving bed sediments or suspended solids
than hyporheic biofilms largely contributing to ER.
Chlorophyll a concentration was also a significant
predictor (r2 ¼ 0.24, P < 0.01) of GPP in Fuirosos.
Chlorophyll responded to the high PAR values in
spring by showing a peak, and these changes had their
effect on stream metabolism. PAR explained 64% of
the variation in GPP and 54% of the variation in chl a
during spring. At the end of April, the canopy closure
caused a 78% reduction in PAR in Fuirosos, which was
accompanied by a remarkable decrease in GPP.
Therefore, it can be concluded that the phenology of
riparian vegetation was an important determinant of
GPP, as stressed by Mulholland et al. (2001). Rosenfeld
& Roff (1991) also reported that light availability, as
mediated by the development of riparian vegetation,
was the most important factor regulating GPP.
Furthermore, only during some weeks in spring were
PAR values higher than Ik, while during most of the
year they were lower, indicating that photosynthesis
was generally light limited (Hill & Dimick, 2002) in
Fuirosos. PAR reaching the streambed after canopy
closure was usually much lower than Ik values, a result
consistent with experimental evidence of light limita-
tion under full canopy (Steinman, 1992; Hill et al.,
1995). These changes in stream metabolism because of
light availability did not have the same effect on ER, in
spite of the increase in chl a concentration. Further-
more, the correlation between chl a and ER was
negative (r ¼ )0.36; P < 0.05), suggesting that break-
down of allochthonous material is driving ER rather
than autotrophic respiration.
Seasonal variations in light and temperature were
not only relevant at an annual scale, but also affected
the daily response of benthic communities. Three
patterns could be differentiated in the P–I curves
obtained in Fuirosos (Fig. 4). GPP was a linear or
quasi-linear function of PAR in late spring, summer
and early autumn. This linear relationship (Fig. 4a)
disappeared in spring with increasing PAR. However,
light saturation of GPP rates was only observed in
April. During this period, Fuirosos GPP became light
saturated when PAR exceeded 90 lE m)2 s)1 (Fig. 4b).
These values are relatively low when compared with
other systems. Hill et al. (1995) gathered data from
multiple studies and showed that photosaturation
typically occurred between 200 and 400 lE m)2 s)1. In
addition, Young & Huryn (1996) provided the first
ecosystem level evidence for light saturation of GPP,
which usually occurred at light intensities over
250 lE m)2 s)1. However, saturation irradiance values
around 100 lE m)2 s)1 are not uncommon in shaded
Mediterranean streams (Guasch & Sabater, 1995). The
hysteresis of the GPP rates (Fig. 4c) observed in late
autumn and winter may be explained by the water
temperature, which, in contrast to PAR, distinctly
differed between morning and afternoon (e.g. by 7 �C).
Ecosystem respiration
Fuirosos is a heterotrophic system because of the
predominance of respiration processes for most of the
year. The P : R ratio averaged 0.16 and annual net
ecosystem production was )1415 g O2 m)2 year)1.
Most of the ER values were reasonably similar to
analogous systems (Table 3). However, ER in autumn
was higher than values recorded previously for
similar forested streams (Table 3). However, there
have been reported by far higher ER values in Prairie
streams, which raise 177 g O2 m)2 year)1 (Wiley,
Osborne & Larimore, 1990). Duffer & Dorris (1966)
reported similar values from North American grass-
land river systems.
The main reason for this extremely high ER in
autumn was the BOM stored in the stream during
summer, which resulted in a high respiration at some
weeks after the flow onset (Fig. 2). Young & Huryn
(1999) also concluded that organic matter supply,
together with light availability, had the most dramatic
influence on community metabolism. Conversely,
Mulholland et al. (2001) determined, in a comparison
of eight streams from several biomes in U.S.A., that
benthic detritus standing crop accounted for only 17%
of ER variation. Also, Webster et al. (1995) reported
that evidence for a positive effect of BOM storage on
ER in eastern U.S.A. streams was weak. The large
BOM accumulation in Fuirosos during low flows
emphasises that its influence on ER is variable not
only on a seasonal basis, but also on an inter-annual
basis. In Fuirosos, one year was drier and allowed
greater storage of BOM, but the following year flow
was sufficient to move BOM downstream.
968 V. Acuna et al.
� 2004 Blackwell Publishing Ltd, Freshwater Biology, 49, 960–971
The second best single predictor of ER in Fuirosos
was water temperature but the relationship between
ER and temperature observed was negative (r2 ¼ 0.37,
P < 0.01). In the Fuirosos, maximum temperatures
were recorded in summer, but this period had
minimum ER values. From a different perspective,
Edwards & Meyer (1987) reported that microbial
respiration during cooler months was stimulated by
large inputs of dissolved and particulate organic
substrates, helping to explain high winter respiration
rates. Therefore, it appears that temperature does not
universally regulate respiration on an annual scale in
stream ecosystems, but other factors (such as BOM
supply and quality, Sinsabaugh, 1997) may also
account for the observed negative relationship
between temperature and ER. In the case of Mediter-
ranean streams, overall climatic variability has to be
considered. Variability in rainfall, which is the prin-
cipal attribute of the Mediterranean-type climate
(Gasith & Resh, 1999), can alter the patterns of the
leaf fall because of drought stress during spring and
summer and, of course, determine the discharge
regime. Therefore, variability in rainfall may have a
major impact on stream metabolism.
In the drier period of this study, the amount of
BOM in Fuirosos was relatively high and steady for
about 3 months. This pattern may be explained by the
enhanced organic matter retention in the stream
channel during low flow (Bilby & Likens, 1980;
Speaker, Moore & Gregory, 1984). However, in
autumn 2002, leaf fall was concentrated in November
and, because of the continuous discharge during this
period, there was not the same BOM accumulation as
recorded the year before. As a consequence, the
values of ER were much higher in autumn 2001 than
in autumn 2002 (Fig. 2). Discharge could also be
controlling a second peak of ER in the middle of
March 2002, related to the remarkable amount of
BOM accumulated during the low winter flow. This
stable discharge period abruptly finished at the end of
March, the BOM decreased to minimum values and
ER also reached minimum values. It is worth empha-
sising that the ER peaks had a certain time lag with
respect to the peaks in BOM (Fig. 2), reflecting the
particular dynamics of leaf detritus processing
(Webster et al., 1995; Schade & Fisher, 1997).
The stream metabolism in Fuirosos seems to
depend on the hydrology and riparian forest inputs,
as they affected both the functioning of primary
producers as well as BOM accumulation and its
respiration. The hydrological regime was mainly
responsible for temporal variation in BOM, and
consequently ER. The rate and timing of entry of
organic matter to stream are major determinants of
stream metabolism.
The lowest stream metabolism activities were
reported in summer, and further research will be
needed to elucidate the importance of pool isolation
during this period. Data from Fuirosos show that there
can be high interannual variability in ER, and that this
is mainly related to hydrological variations between
years. Although these patterns may not be unique to
Mediterranean-type systems (e.g. Uehlinger, 2000), it
is obvious in these systems that the influence of flow
extremes, such as spates or extended periods of low
flow, may be able to modify the relative contributions
of both GPP and ER to lotic metabolism.
Acknowledgments
Andrea Butturini, Nuria Morral, Anna Romanı, Joan
Artigas, Elena Guerra, Meritxell Aznar and Ainhoa
Gaudes assisted in the stream and in the laboratory.
Maurice Lock (University of Wales) and two anony-
mous reviewers provided useful comments on the
manuscript. This research was funded by the CICYT
projects AMB99-0499 and REN2002-04442-C02-02/
GLO of the Spanish Science Ministry.
References
APHA (1992) Standard Methods for the Examination of
Water and Wastewater, 18th edn. Washington DC,
U.S.A.
Bernal S., Butturini A. & Sabater F. (2002) Variability of
DOC and nitrate responses to storms in a small
Mediterranean forested catchment. Hydrology and Earth
System Sciences, 6, 1031–1041.
Bilby R.E. & Likens G.E. (1980) Importance of organic
debris dams in the structure and function of stream
ecosystems. Ecology, 61, 1107–1113.
Bott T.L., Brock J.T., Dunn C.S., Naiman R.J., Ovink R.W.
& Petersen R.C. (1985) Benthic community metabolism
in four temperate stream systems: an interbiome
comparison and evaluation of the river continuum
concept. Hydrobiologia, 123, 3–45.
Buhrer H. (1975) Computerprogramm zur Bekanntgabe
aktueller Seedaten. Schweizerische Zeitschrift Fur
Hydrologie, 37, 332–346.
Riparian influences on stream metabolism 969
� 2004 Blackwell Publishing Ltd, Freshwater Biology, 49, 960–971
Chessman B.C. (1985) Estimates of ecosystem metabo-
lism in La Trobe River, Victoria. Australian Journal of
Marine and Freshwater Research, 36, 1354–1364.
Cummins K.W., Wilzbach M.A., Gates D.M., Perry J.B. &
Taliaferro W.B. (1989) Shredders and riparian vegeta-
tion: leaf litter that falls into streams influences
communities of stream invertebrates. Bioscience, 39,
24–30.
Duffer W.R. & Dorris T.C. (1966) Primary Productivity in
a southern Great Plains Stream. Limnology and Oceano-
graphy, 11, 143–151.
Edwards R.W. & Meyer J.L. (1987) Metabolism of a sub-
tropical low gradient blackwater river. Freshwater
Biology, 17, 251–263.
Edwards R.W. & Owens M. (1962) The effects of plants
on river conditions. IV. The oxygen balance of a chalk
stream. Journal of Ecology, 50, 207–220.
Fellows C.S., Valett H.M. & Dahm C.N. (2001) Whole-
stream metabolism in two montane streams:
Contribution of the hyporheic zone. Limnology and
Oceanography, 46, 523–531.
Fisher S.G. & Carpenter S.R. (1976) Ecosystem and
macrophyte primary productivity of the Fort River,
Massachusetts. Hydrobiologia, 47, 175–187.
Gasith A. & Resh V.H. (1999) Streams in Mediterranean
Climate Regions: Abiotic Influences and Biotic
Responses to Predictable Seasonal Events. Annual
Review of Ecology and Systematics, 30, 51–81.
Genereux D.P. & Hemond H.F. (1992) Determination of
Gas Exchange Rate Constants for a Small Stream on
Walker Branch Watershed, Tenessee. Water Resources
Research, 28, 2365–2374.
Gordon N.D., McMahon T.A. & Finlayson B.L. (1992)
Stream Hydrology. An Introduction for Ecologists. Wiley,
Cichester, United Kingdom.
Gregory S.G., Swanson F.J., Mckree W.A. & Cummins
K.W. (1991) An ecosystem perspective of riparian
zones. Bioscience, 41, 540–551.
Guasch H. & Sabater S. (1995) Seasonal variations in
photosynthesis-irradiance responses by biofilms in
Mediterranean streams. Journal of Phycology, 31, 727–
735.
Hall C.A.S. (1972) Migration and metabolism in a
temperate stream ecosystem. Ecology, 53, 585–604.
Henley W.J. (1993) Measurement and interpretation of
photosynthetic light-response curves in algae in the
context of photoinhibition and diel changes. Journal of
Phycology, 29, 729–739.
Hill W.R. & Dimick S.M. (2002) Effects of riparian
leaf dynamics on periphyton photosynthesis and light
utilisation efficiency. Freshwater Biology, 47, 1245–
1256.
Hill W.R., Ryon M.G. & Schilling E.M. (1995) Light
limitation in a stream ecosystem: responses by primary
producers and consumers. Ecology, 76, 1297–1309.
Hill W.R., Mulholland P.J. & Marzolf E.R. (2001) Stream
Ecosystem Responses to Forest Leaf Emergence in
Spring. Ecology, 82, 2306–2319.
Hoskin C.M. (1959) Studies of Oxygen Metabolism of
Streams of North Carolina, Vol. 6. Publications of the
Institute of Marine Science. Published by the Institute
of Marine Science, The University of Texas, Port
Aranas, Texas, pp. 186–192.
Jassby A.D. & Platt T. (1976) Mathematical formulation of
the relationship between photosynthesis and light for
phytoplankton. Limnology and Oceanography, 21, 540–
547.
Jeffrey S.W. & Humphrey G.F. (1975) New spectro-
photometric equations for determining chlorophylls
a,b,c and c2 in higher plants, algae and natural
phytoplankton. Biochemie und Physiologie der Pflanzen,
167, 191–194.
Kaenel B.R., Buehrer H. & Uehlinger U. (2000) Effects of
aquatic plant management on stream metabolism and
oxygen balance in streams. Freshwater Biology, 45, 85–
95.
Lamberti G.A. & Steinman A.D. (1997) A comparison of
primary production in stream ecosystems. Journal of the
North American Benthological Society, 16, 95–104.
Margalef R. (1983) Limnologıa. Omega, Barcelona, Spain.
Marzolf E.R., Mulholland P.J. & Steinman A.D. (1998)
Reply: improvements to the diurnal upstream-down-
stream dissolved oxygen change technique for deter-
mining whole-stream metabolism in small streams.
Canadian Journal of Fisheries and Aquatic Sciences, 55,
1786–1787.
McCree K.J. (1971) Test of current definitions of Photo-
synthetically Active Radiation against leaf photosyn-
thesis data. Agricultural Meteorology, 10, 443–453.
Minshall W.G. (1978) Autotrophy in stream ecosystems.
Bioscience, 28, 767–771.
Molla S., Maltchik L., Casado C. & Montes C. (1996)
Particulate organic matter and ecosystem metabolism
dynamics in a temporary Mediterranean stream.
Archive fur Hydrobiologie, 137, 59–76.
Mulholland P.J., Fellows C.S., Tank J.L. et al. (2001) Inter-
biome comparison of factors controlling stream meta-
bolism. Freshwater Biology, 46, 1503–1517.
Odum H.T. (1956) Primary production in flowing waters.
Limnology and Oceanography, 1, 102–117.
Rathbun R.E., Stephens D.W., Schultz D.J. & Tai D.Y.
(1978) Laboratory studies of gas tracers for reaeration.
Journal of Environmental Engineering Proc ASCE, 104,
215–229.
970 V. Acuna et al.
� 2004 Blackwell Publishing Ltd, Freshwater Biology, 49, 960–971
Rosenfeld J. & Roff J.C. (1991) Primary production and
potential availability of autochthonous carbon in
southern Ontario streams. Hydrobiologia, 224, 99–109.
Sabater S., Bernal S., Butturini A., Nin E. & Sabater F.
(2001) Wood and leaf debris input in a Mediterranean
stream: the influence of riparian vegetation. Archive fur
Hydrobiology, 153, 91–102.
Schade J.D. & Fisher S.G. (1997) Leaf litter in a Sonoran
desert stream ecosystem. Journal of the North American
Benthological Society, 16, 612–626.
Schwarz G. (1978) Estimating the dimension of a model.
The Annals of Statistics, 6, 461–464.
Shoup T.E. (1983) Numerical Methods for the Personal
Computer. Prentice-Hall, Englewood Cliffs, N.J.
Sinsabaugh R.L. (1997) Large-scale trends for stream
benthic respiration. Journal of the North American
Benthological Society, 16, 119–122.
Smock L.A., Metzler G.M. & Gladden J.E. (1989) Role of
debris dams in the structure and functioning of low-
gradient headwater streams. Ecology, 70, 764–775.
Speaker R., Moore K. & Gregory S. (1984) Analysis of the
process of retention of organic matter in stream
ecosystems. Verhandlungen der Internationale Vereini-
gung fur Theoretische und Angewandte Limnologie, 22,
1835–1841.
Steinman A.D. (1992) Does an increase in irradiance
influence periphyton in heavily-grazed woodland
stream? Oecologia, 91, 163–170.
Stream Solute Workshop (1990) Concepts and methods
for assessing solute dynamics in stream ecosystems.
Journal of the North American Benthological Society, 9, 95–
119.
Tsivoglou E.C. & Neal L.A. (1976) Tracer measurement of
reaeration III. predicting the reaeration capacity of
inland streams. Journal of Water Pollution Control
Federation, 48, 2669–2689.
Uehlinger U. (2000) Resistance and resilience of ecosys-
tem metabolism in a flood-prone river system. Fresh-
water Biology, 45, 319–332.
Uehlinger U. & Naegeli M.W. (1998) Ecosystem metabo-
lism, disturbance, and stability in a prealpine gravel
bed river. Journal of the North American Benthological
Society, 17, 165–178.
Uehlinger U., Konig C. & Reichert P. (2000) Variability of
photosynthesis-irradiance curves and ecosystem re-
spiration in a small river. Freshwater Biology, 44, 493–
507.
Uehlinger U., Naegeli M.W. & Fisher S.G. (2002) A
heterotrophic desert stream? The role of sediment
stability. Western North American Naturalist, 62, 433–
473.
Vannote R.L., Minshall G.W., Cummins K.W., Sedell J.R.
& Cushing C.E. (1980) The River Continuum Concept.
Canadian Journal of Fisheries and Aquatic Sciences, 37,
130–137.
Webster J.R., Wallace J.B. & Benfield E.F. (1995) Organic
processes in streams of the eastern United States.
In: River and Streams Ecosystems (Eds C.E. Cushing,
K.W. Cummins & G.W. Minshall), pp. 117–187. Else-
vier Science, Amsterdam, The Netherlands.
Wiley M.J., Osborne L.L. & Larimore R.W. (1990) Long-
itudinal structure of an agricultural prairie river
system and its relationship to current stream ecosys-
tem theory. Canadian Journal of Fisheries and Aquatic
Sciences, 47, 373–384.
Young R.G. & Huryn A.D. (1996) Interannual variation in
discharge controls ecosystem metabolism along a
grassland river continuum. Canadian Journal of Fisheries
and Aquatic Sciences, 53, 2199–2211.
Young R.G. & Huryn A.D. (1999) Effects of land use on
stream metabolism and organic matter turnover.
Ecological Applocations, 9, 1359–1376.
(Manuscript accepted 23 April 2004)
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