International Journal of Salt Lake Research 7: 187–210, 1998.
© 1998 Kluwer Academic Publishers. Printed in the Netherlands.
Anthropogenic perturbations to the trophic structure
in a permanent hypersaline shallow lake: La Salada de
Chiprana (north-eastern Spain)
P. DÍAZ1,3, M.C. GUERRERO1, P. ALCORLO1, A. BALTANÁS1, M.
FLORÍN1,2 and C. MONTES1
1 Departamento de Ecología de la Universidad Autónoma de Madrid, 28049 Cantoblanco,
Madrid, Spain; 2 Present address: DPT. Plant. Ecol. Evolut. Biol., Utrecht University, P.O.
Box 80084, 3508 TB Utrecht, The Netherlands; 3 Present address: Dpto. de Ciencias
Ecológicas, Facultad de Ciencias, Universidad de Chile. P.O. Box 653, Santiago de Chile
Abstract. The changes in the trophic state in the Salada de Chiprana (north-eastern Spain)
over two quite different seasonal cycles (1989, 1994/95) were studied. During the former
cycle, the lake was permanently stratified, and was biogenically meromictic, and in the latter,
showed no apparent stratification. The main variables related to the physico-chemical changes
observed can be attributed to the effect caused by the increase in the nutrient loading. The
large amounts of nutrients (total-N and total-P) and organic matter are due to the use of
the lake as a reservoir for water discharged from irrigation. Two remarkable effects of the
change are the permanent mixing of the water column and the immobilization of phosphorus
in the form of ionic species and solid phases that are not available to the biota especially
primary producers (phytoplankton, periphyton, microbial mats). The results of the present
study emphasize the fragility of (hyper) saline ecosystems to anthropogenic disturbances such
as increases in freshwater inflow and nutrient inputs. Likewise, the study reveals the failure of
conservation criteria that have been used to manage this lake, especially those which refer to
the control of freshwater, nutrient-rich influents.
Key words: microbial mats, mixis, nutrient loading, periphyton, saline lakes
Introduction
The study of saline lakes is often considered an unusual branch of limnology,
but this is because these environments are rare in most temperate countries.
In Spain, saline lakes are a common feature of the landscape in arid and
semi-arid regions (Montes and Martino, 1987; Alonso, 1987) where they
provide unique opportunities for ecological studies of both the structure and
functioning of such ecosystems (Williams, 1993).
Most saline lakes of continental origin are found in endorheic areas which
are subject to extreme climatic conditions, sometimes aggravated by agricul-
188
tural practices or other human activities in areas where drought is common. In
Spain, approximately 70 lakes are situated in arid and semi-arid zones. They
occur in four large endorheic regions in the Iberian peninsula: the Duero, La
Mancha, Andalucía and Aragón catchments (Martino et al., 1986a).
Generally, Spanish saline lakes are relatively well conserved compared
to freshwater and sub-saline ecosystems (Martino et al., 1986 b). However,
like some of the largest saline lakes of the world (Lake Mono, Jellison and
Melack, 1988; Pyramid Lake, Reuter et al., 1993; Hot Lake, Anderson, 1958,
U.S.A.), the smaller saline lakes in Spain (Lake La Sariñena, Huesca, Martino
et al., 1986b) frequently receive inflows of fresh water enriched by nutrients
from anthropogenic activities (e.g. agricultural practices) in the catchments
(Martino et al., 1986b; Williams, 1993). In many cases, this results in degradation of the ecosystem through eutrophication (Hammer, 1978). These
activities usually occur over a long time and few studies were carried out
before the lakes were affected. Accordingly, it is difficult to describe trophic
changes and other responses of these saline ecosystems to perturbation.
Functional studies of hypersaline lakes with a dominant sulphate-magnesium
character are particularly rare (Robarts et al., 1992). It has also been observed
that the physico-chemical complexity of these particular lakes strongly affects
nutrient availability; this results in a different metabolic pattern in response
to changes in trophic state and/or salinity to that found in fresh water (Caraco
et al., 1989).
To understand these phenomena better, there is a need to carry out studies which (1) monitor complete hydrological cycles in saline environments,
and (2) relate environmental factors to the mechanisms of biological control,
rather than the more usual studies which are restricted to descriptions of daily
fluctuations, food chains and species lists (Jellinson and Melack, 1988). A
combination of such studies with those describing daily fluctuations would
improve knowledge about saline lakes by allowing the recognition of new criteria for ecological assessment which could be used for the conservation and
protection of these ecosystems. Often they are threatened through ignorance
of the key processes that underpin their scientific, economic and cultural
value. Knowledge of the mechanisms controlling the functioning of these
lakes at different scales will avoid the application of negative management
practices, such as an increased freshwater supply to increase the period of
inundation (Williams, 1981 and 1993; Martino et al., 1986b).
The objectives of this study were: (1) to analyse both the changes in the
trophic trend of a hypersaline lake subjected to extreme anthropogenic perturbations, and the effect of these changes on the structure and dynamics of
the community of primary producers, and (2) to compare the results obtained
189
from a hypersaline lake with models used in freshwater ecosystems equally
perturbed by an excessive increase in nutrient supply.
Study area
The Salada de Chiprana (41◦ 140 N; 0◦ 100 W) is the deepest permanent hypersaline lake (Z max. = 5 m) in the Iberian Peninsula and was included in the
Ramsar Convention in 1994 because of its singular ecological value. It is
located in the central eastern zone of the River Ebro basin, in the Region of
Bajo Aragón (Zaragoza), approximately 5 km far from the village of Chiprana
(Figure 1). The lake is located at a lower altitude than any other lake within
the Ebro catchment (150m a.s.l.). The climate is semi-arid, with a strong
water deficit aggravated by intense dry winds (Guerrero et al., 1991).
Geologically, the Salada de Chiprana lies on detritic material of unusual
litho-morphological character (Riba et al., 1967; Friend et al., 1986). The bed
of the lake comprises an ancient network of Tertiary paleo-channels composed of fossilized sands which now remain as ridges due to the differential
erosion of the less resistant silt deposits between channels. As surface water
drainage is poorly developed, the flow of groundwater is the most important
factor in the maintenance of water within the lake.
The lake forms part of a wetland complex which includes three other lakes:
L. Salobrosa, L. Prado del Farol and La Estanca de Chiprana. The latter is the
only one which presently contains water. This is because it has been converted
into a reservoir to impound water from the Guadalope River. The other two
have dried out from the accumulation of sediments and are covered with
dense reedbeds. The principal cause of this process is the inflow of drainage
water with high sediment loads from the irrigated lands that surround the
lakes.
The Salada de Chiprana has also experienced inflows of fresh water since
the 1980’s, when a canal was excavated to connect it to Lake Salobrosa in
order to drain excess irrigation water flowing into the latter (Berga et al.,
1994). After 1991, the connecting canal, which was covered by reed (Phragmites spp.), was cemented. This subsequently eliminated nutrient retention
by plants. Since then, fresh water with high concentrations of nutrients and
probably pesticides directly enters into the lake, with the highest discharges
occurring between March and July. Furthermore, dry groves of olive-trees
surrounding the lake were partially replaced by irrigation crops (cereal,
maize, alfalfa) causing additional freshwater inputs to the lake as diffuse
runnoff (Berga et al., 1994).
According to data obtained during the 1989 hydrological cycle (Guerrero et al., 1991; Vidondo et al., 1993), together with other isolated records
190
Figure 1. Location of Salada de Chiprana in the Iberian Peninsula and locally. Bathymetric
map of Salada, showing sampling station (∗) of deepest zone (1) and shore one (2), freshwater
inlet (→), submersed paleo-channels ( ), and marginal vegetation distribution ( ).
191
(Martino, 1988; Montes, unpublished data), average salinity varies between
30 and 73 g L−1 ; the Salada de Chiprana may thus be classified as a αhypersaline lake (Por, 1980). Its water is of sulphate-magnesium character
due to the evaporation of salts (mainly gypsum and dolomite). During the
1989 hydrological cycle, the sediments of the lake were almost permanently
anaerobic and contained high concentrations of H2 S (Guerrero et al., 1991;
Vidondo et al., 1993).
Under these extreme environmental conditions, only a limited number of
species is able to survive. Most communities are planktonic and microbenthonic. Extensive areas of emergent vegetation occur in the shallow areas.
An interesting feature is the presence of microbial mats composed of benthic
communities of micro-organisms dominated by cyanobacteria (Jørgensen et
al., 1983). These mats are formed by the entrapment, cohesion and precipitation of sediment particles following the growth and metabolic activity of
micro-organisms (Walter, 1976). In the case of Salada de Chiprana, microbial
mats contain diatoms and cyanobacteria, and they cover the sediments of
shallow shores of moderate slope up to a depth of 1.5 m.
Both the zooplankton and the aquatic vegetation are represented by
species adapted to high salinity as indicated by the presence of the Anostraca Artemia parthenogenetica (Amat, 1995), characteristic of permanent
saline environments, and the presence of macrophytes such as Ruppia maritima L. var. maritima, R. drepanensis Tineo, and charophytes such as
Lamprothamnium papulosum.
The faunal community is characterised by low species diversity involving
several cosmopolitan halophilous taxa: Fabrea salina (Ciliata), Hexarthra
fennica Schmarda and Brachionus plicatilis (O.F. Müller) (Rotifera), Arctodiaptomus salinus (Daday) and Cletocamptus retrogressus Schmankewitsch
(Copepoda) and Ephydra spp. (Diptera).
Materials and methods
Information on the hydrological cycle 1988–89 corresponds to that of previous studies of our group (Vidondo, 1991; Vidondo and Guerrero, 1992;
Vidondo et al., 1993).
Monthly sampling for the hydrological cycle 1994/95 was carried out
between November and July. Two sampling stations in the lake were selected
(Figure 1): The first (Sta. 1) was located in one of the deepest areas (Zmax.
= 5m), and the second one (Sta. 2) in shallow water in one of the flat beach
areas.
Temperature, conductivity, pH, light penetration, and the concentration of
oxygen, alkalinity and hydrogen sulphide were measured in situ at various
192
depths (each 0.5 m from surface to bottom) only at station 1. Conductivity and
temperature were measured with a conductivity meter (WTW LF96), pH was
measured with a portable pH meter (Crison 506), and dissolved oxygen was
measured using an oxygen meter (WTW EOT-196). Alkalinity and hydrogen
sulphide were measured in situ using the volumetric method of Aminot and
Chaussepied (1983) and a Merck hydrogen sulphide kit, respectively.
Water samples from station 1 were collected and stored in 1 L glass
bottles, pre-washed with 5% HCl. They were analysed for total N and P,
alkalinity, total dissolved solids, and chlorophyll a. When necessary, samples were stored on ice and in darkness for subsequent laboratory analysis.
During 1994/95, in the area of the lake influenced by the freshwater inflow
from the drainage canal, total-P concentration was recorded from the point of
entry into the lake (Figure 1) to a distance of approximately 150 m along the
shoreline.
At station 1, phytoplankton samples were collected in 125 ml dark glass
bottles and fixed in situ with Lugol’s solution and samples for the measurement of chlorophyll a concentration were passed through GF/C glass fibre
filters.
At station 2, samples of microbial mats and macrophytes were collected
with plexiglas cores (45 mm diameter). The samples were fixed in 4 per cent
formalin for subsequent identification in the laboratory.
The planktonic and benthic fauna were sampled from June 1994 until
October 1995. In the shallow water area, zooplankton was sampled qualitatively in shore zones using a hand-net (60 µm mesh). In the deep water area
(station 1), zooplankton was sampled quantitatively by filtering 10 L of water
at each depth. Five replicates were taken at one metre depth intervals from
the surface to the bottom of the lake. Zoobenthos was sampled quantitatively
by taking two replicates at three points along a transect between station 1 to
station 2, using a Mondsee core and fixing samples in the field with 4 per cent
neutralised formalin.
In the laboratory, total N and P were measured using non-filtered water
samples and the persulphate digestion technique (APHA, 1992). Total dissolved solids were measured by desiccation at 105 ◦ C (APHA, 1992). Chlorophyll a was extracted in 90 per cent acetone and measured using the technique
of Jeffrey and Humphrey (1975).
Phytoplankton was sedimented-out in Utermöhl boxes and examined with
an inverted microscope (Olympus, model CK2). Microbial mats and macrophytes were studied using a stereo and compound microscope (Olympus,
model BH-2). Zooplankton and zoobenthos were preserved in 70% ethanol
and observed with a stereo (SZ-30) and compound microscope (Olympus,
model BH-2) when necessary.
193
Table 1. Discrete values of surface biotic and abiotic variables in the Salada de
Chiprana in February 1986, March 1987 and May 1988 (Martino, 1988; Montes,
unpublished data).
Variable
February 1986
March 1987
May 1998
Conductivity (mS cm−1 )
Oxygen (mg L−1 )
Chlorophyll a a.z. (µg L−1 )
Chlorophyll a an.z. (µg L−1 )
Total-P (µg-at L−1 )
Total-N (µg-at L−1 )
55
8.4
2.01
—
3.58
175
53.3
10.8
8
—
2.9
347.4
41.9
7.15
1.65
220.9
—
74.57
The physico-chemical balance and ionic speciation were determined using
the computer programme WATEQ (Plummer et al., 1984).
Results
Table 1 presents previously unpublished data for 1986, 1987 and 1988 for the
Salada de Chiprana. Table 2 shows mean values, variance and range of the
most important biotic and abiotic variables measured seasonally during the
hydrological cycles of 1988–89 (Vidondo, 1991) and 1994/95.
Changes in environmental variables and the stratification pattern in the
water column
With regard to water-level fluctuations, 1989 showed greater stability with an
annual variation of 60 cm compared with the period 1994/95 when variation
was 1 m. 1994/95 was also significantly drier than 1989 (Díaz, 1998).
Both conductivity and salinity are notably lower in 1989 compared with
other periods, with a minimum of 27.2 mS cm−1 and 29.6 g L−1 at the surface
and a maximum of 51.8 mS cm−1 and 70.9 g L−1 on the lake bed, respectively
(Tables 1 and 2). The values for the period 1994/95 show an increase in conductivity (Table 2). Dissolved oxygen concentration was, on average, lower
in 1989 than in the other two hydrological cycles, due to the presence of an
anoxic layer during most of that year. Oxygen concentration ranged from 0
to 7.92 mg L−1 in 1989 and 0.1 to 15.6 mg L−1 in 1994/95.
The vertical profiles of the most important abiotic variables are shown in
Figure 2. Three seasonal periods in 1989 (Vidondo, 1991) were selected for
direct comparison of the mixing and/or stratification of the water column with
the equivalent seasonal period in 1994/95.
Season
Autumn
X∗
S∗
Range
Winter
X∗
S∗
Spring
X∗
Range
S∗
Range
Summer
X∗
S∗
Range
1989
Temperature (◦ C)
17.3
0.23
0.00–17
11
1.41
9.20–13.6
15.6
2.23 15.30–22.8
23.18
Conductivity (mS cm−1 )
51.4
0.24 51.10–51.8
47.8
5.47 27.20–51.2
44.75
2.56 41.60–50
38.79
55.27
—
—
53.2
—
—
63.2
—
—
59.96
T.D.S.∗ (g L−1 )
Oxygen (mg L−1 )
2.3
0.44
0.00–2.65
4
2.74
0.00–7.92
3.96
1.54
0.00–6.17
5.43
Chlorophyll a a.z. (µg L−1 )
86.6
49.95 25.00–144
3.9
8.96
4.57–34
14.67
8.69
0.93–27.4
8.44
25.00–4743.3 173.1 168.3
25.40–385.5 2161.28 1903.8
69.90–2206.8 848.58
Chlorophyll a an.z. (µg L−1 ) 2262.5 2461.6
Total-P (µg-at L−1 )
<0.03
—
<0.03–0.94
<0.03
—
<0.03–0.11
<0.03
—
<0.03–0.64
0.14
Total-N(µg-at L−1 )
15.47
23.06
1.32–74.61
6.5
4.05
2.77–12.81
20.86
11.2
3.22–58.89
3.00
2.1
18.50–24
5.08
37.70–48.6
—
—
2.09
0.00–6.61
7.47
5.84–45
717.83
32.40–4616.2
0.209 <0.03–0.51
1.67
10.95
1994/95
Temperature (◦ C)
Conductivity (mS cm−1 )
T.D.S.∗ (g L−1 )
Oxygen (mg L−1 )
Chlorophyll a a.z. (µg L−1 )
Chlorophyll a an.z. (µg L−1 )
Total-P (µg-at L−1 )
Total-N(µg-at L−1 )
18.54
58.95
80.42
3.62
3.79
—
0.05
449
1.98 15.90–20.1
2.2
57.00–61.1
—
—
4.05
0.30–8.7
2.82 <0.01–4.93
—
—
1.25
1.03–1.5
428.9 398.20–539.6
6.58
54.27
83.69
6.36
21.74
—
0.63
590.8
∗ TDS: Total disolved solids, X: Means, S: Standard deviation.
0.63
5.90–7.8
0.4
53.30–60.5
—
—
1.59
3.70–10.1
47.23
0.47–128.7
—
—
0.16
0.27–12.21
57.31 576.30–713.9
11.67
0.18 13.10–20.3
54.38
0.18 54.10–62.1
85.61
—
92.22
13.39
0.34
8.60–15.6
5.7
7.22
2.23–8.05
—
—
—
14.70
51.9
0.31–22.2
601.89 1955.7 521.30–644.3
27.53
0.8
47.42
0.76
—
—
3.65
3.1
4.49
4.3
—
—
3.075
0.28
442.5
259.7
25.90–28.3
46.20–63.2
0.10–7.8
2.23–8.06
—
3.02–3.7
379.00–669.7
194
Table 2. Mean values and variation range of main biotic and abiotic parameters recorded in the deepest zone (station 2) of the Salada de Chiprana during
two hydrological periods: 1989 (Vidondo, 1991) and 1994/95.
195
Figure 2. Comparison of abiotic variables in 1989 and 1994/95. Circles represent conductivity of water, crosses represent temperature, suns represent
dissolved oxygen concentration, and full line represents concentration of hydrogen sulphide. The shaded area represents the anaerobic zone.
196
The annual cycle of 1989 is characterised by the presence of a long period
of meromixis (with anoxia in the deeper layers) due to biogenic causes, with
the exception of October, when complete mixing of the water column took
place (Figure 2A) and the maximum value of dissolved oxygen on the bed
of the lake was recorded (1.3 mg L−1 ). In other months, the deepest part of
the lake was without oxygen. Comparison in the autumn of the two hydrological cycles (Figures 2A and 2A1) showed that, whereas in 1989 the water
column was completely mixed, in 1994 the lake showed a definite oxycline,
though without a complete depletion of oxygen. Likewise, a thermocline at
2.5 metres depth was observed and the conductivity showed a slight increase
with depth.
Figures 2B and 2B1 are also quite different, with a clear oxycline and deep
anaerobic bottom in January 1989, compared with a situation of mixing and
only slight depletion of oxygen with depth in 1995. Temperature increased
slightly with depth in 1989, as did conductivity, whereas in 1995 the tendency
was nearly the opposite. In winter, conductivity remained fairly uniform at all
depths.
Finally, in July (Figures 2C and 2C1), the profiles for 1989 and 1994/95
were rather similar and showed a total depletion of oxygen below 4 m depth.
Temperature and conductivity remained similar for both cycles, with the former decreasing slightly with depth, and the latter showing a slight increase.
Figure 2C shows high levels of hydrogen sulphide in the anoxic zone, while
in July 1995 (Figure 2C1), the values were almost at the limit of detection.
Changes in the nutrient loading, chlorophyll concentration and anoxic
conditions
Data shown in Tables 1 and 2 reveal that nutrients are the parameters that
generally varied most significantly through time ( Student’s t = –5.4 and –
35.19 for P and N, respectively; p < 0.05). Figure 3 shows the variation in
phosphorus and nitrogen values for the different sampling periods: maxima
for both phosphorus (spring) and nitrogen (winter) during the cycle 1994/95
were 500 times greater than those measured in 1989 at the same periods.
Values for both nutrients and chlorophyll a in 1986 and 1987 are similar to
those of 1994/95.
The increase in the loading of nutrients observed between 1989 and
1994/95 generated a change in the trophic state of the lake from oligotrophic
in 1989 to mesotrophic in 1994/95. However, analysis of ionic speciation
and chemical equilibrium in regard to phosphorus (Plummer et al., 1984) in
the Salada during 1994/95 shows that the percentage of orthophosphate in
relation to total P was less than 8 per cent and most P was found in the form
of tightly bound chemical complexes and unavailable for use by planktonic
197
Figure 3. Seasonal variation in the mean values of total-P and total-N from 1986–1989 (Martino, 1988; Vidondo, 1991; Montes, unpublished data) and 1994/95. I: autumm; II: winter;
III: spring; and IV: summer. Maximum, minimum and median values are indicated on the left
axis, and mean values on the right axis.
198
Figure 4. Ionic speciation and physico-chemical equilibrium of phosphorus to 1994/95.
primary producers (Figure 4). The availability of other chemical forms such
as HPO4 and H2 PO4 could depend on the range of pH variation. Of the
unavailable P, 44% was found in combination with magnesium.
Figure 5 shows the relationship of seasonal mean concentrations of N and
P in 1989 and 1994/95, with the line representing the minimum N:P ratio of
15:1. The N values for 1994/95 are always higher than those for 1989 and
above the regression line, indicating P-limitation. On the other hand, values
for 1989 lie in the central-left part of the graph, close to the regression line,
with alternating limitations by N and P, but a clear predominance of the latter.
If it is considered that only 8 per cent of the total-P is available to the biota, it
might be expected that values of 1994/95 would move further away from the
axis.
Related to the effect of the drainage canal into the Salada, the seasonal
dynamics of total-P in maximum freshwater inflows (Figure 1) show a narrow
correlation with those in the surface layer of the Salada (r: 0.7, p < 0.05).
Figure 6 shows the vertical profiles of the means and standard deviations
of chlorophyll a during the two periods, 1989 and 1994/95. In 1989, values were quite uniform to a depth of 3 m, after which there was a marked
increase. This increase coincided with the average height of the anaerobic
zone during this study and the vertical limit of green photosynthetic sulphur
199
Figure 5. N/P ratios (15:1). Squares: 1989 (Vidondo, 1991); circles: 1994/95.
bacteria, as documented in other studies (Vidondo et al., 1993). Maximum
values of chlorophyll a in 1989 were always recorded below 4 m, but values
recorded in 1994/5 had a maximum at 3 m, decreasing below this depth. This
change in the pattern of chlorophyll a concentrations in the water column did
not increase in parallel with the change in the nutrient concentrations. In fact,
the opposite trend is observed, with values for 1994/95 on average 85 per
cent lower than those for 1989. Values for chorophyll a in 1989 showed their
maxima in the anaerobic zone of the water column.
In 1989, recorded values of hydrogen sulphide were high (annual mean =
18.88 mg L−1 ) in the lower depths of the lake, which remained anoxic during
most of the hydrological cycle. The 1994/95 period, on the other hand, was
characterised by the almost total absence of hydrogen sulphide and by lower
layers of water containing a substantial amount of oxygen throughout most
of the sampling period except in July. In this period, the maximum value
of hydrogen sulphide was recorded (1 mg L−1 ) on the lake-bed, and anoxic
conditions prevailed.
The presence of oxygen in 1994/95 coincided with the disappearance of
communities of green photosynthetic sulphur bacteria. These utilise hydrogen
sulphide liberated in the sediments as an electron donor in the process of
photosynthesis.
200
Figure 6. Vertical profile of mean values of chlorophyll a in 1989 (Vidondo, 1991) and
1994/95.
Submerged vegetation and microbial mats
Aquatic vegetation showed a marked decline between 1988/89 and 1994/95.
In particular, the extent of Ruppia maritima var. maritima and Lamprothamnium papulosum decreased dramatically. During this investigation, the
distribution of both communities was limited to a few short segments of the
shoreline.
An important observation in 1994/95 was that both macrophyte species
were found to be covered with a dense growth of golden-brown periphyton
during total the hydrological cycle, thus limiting their capacity to capture
light.
201
Figure 7. Schematic representation of vertical section through a summer mat, showing the
different constituent layers. In 1989 (Guerrero et al., 1991) the letters within layers refer to:
a: Diatom spp.; b: Gloeocapsa sp.; c: Synechococcus sp.; d: Microcoleus chthonoplastes; e:
Spirulina sp.; f: Oscillatoria spp.; g: Pseudoanabaena sp.; h: Flexibacteria; i: Chromatium sp.;
j: Thiospirillum sp.; A: green layer, B: red layer, C: sediments. In 1994/95, a: Cymbella aspera;
b: Cyanobacteria ‘LPP’; c: Gloeocapsa sp.; d: Diploneis sp.; e: Synedra sp.; f: Cocconeis sp.;
g: Spirulina sp.; h: Thiospirillum sp.; i: Oscillatoria sp.; SUP: golden-brown layer; A: red
Layer; B: green layer; C: sediments.
Both entire and germinated seeds of Ruppia sp. were found in various
states of decomposition some centimetres below the surface of the sediment. The sediment itself contained considerably more organic material than
recorded in 1989.
Only at the end of summer 1995 was there a slight recovery of Lamprothamnium papulosum. This was restricted to those zones of the shore where a
carpet of younger microbial mats had become established. These areas were
on the surface of the old paleo-channels where sediment contains more sand
and less organic material.
The microbial mats have also suffered changes since 1989: on a macroscopic scale, they are now less compacted. The thickness of the periphyton
surface layer has increased and the surface of the mat is no longer covered
in cone-like projections referred to as a “pinnacle mat” (Gulobíc, 1973). Figure 7 is a diagram showing the principal microscopic changes which have
occurred between the two stages.
Note that although the structural layering of the mats was retained, both
the relative position of the layers and the specific composition within the
202
mat changed considerably. In 1994/95 the most important change in the
uppermost, surface layer (Figure 7, Sup) was the extensive development of
a golden-brown, thick layer of pennate diatoms (Cymbella sp., Synedra sp.,
Nitzschia spp., Pinnularia sp., Navicula sp., Diploneis sp., Cocconeis sp.),
together with cyanobacteria such as Gloeocapsa sp. and ‘LPP’ forms (Rippka
et al., 1979) in lower proportions. In 1989, the equivalent surface layer was
extremely thin and almost indistinct from the immediately adjacent green
thick layer.
The intermediate layer in 1994/95 was composed of a thin reddish layer
(1–2 mm thick). It differed from that of 1989, which was a green layer
(3–8 mm thick) comprised mostly of Microcoleus chthonoplastes (Geitler,
1932) and other filamentous forms which varied in proportion during the
year (Guerrero et al., 1991). The present reddish layer is composed mostly
of Thiospirillum sp. (red photosynthetic sulphur bacteria) and some less
numerous ‘LPP’ forms, Gloeocapsa sp. and Spirulina sp. Another red sulphur bacterium Chromatium, which was present in 1989, was not recorded in
1994/5.
There was an intensely green layer (1–2 mm thick) immediately below
these two layers composed mainly of diatoms (Diploneis sp., Synedra sp.
and Cymbella sp.) and cyanobacteria (mostly ‘LPP’ forms, Gloeocapsa sp.
and Oscillatoria sp.). The green layer has changed in its position since 1989;
in 1989 it was the thickest layer under the uppermost thin superficial one,
while in 1994/95 it was situated in the deepest part of the mat under the red
layer. Also noticeable in 1994/95 was the disappearance of the filamentous
cyanobacterium Microcoleus chthonoplastes as the main constituent of the
green layer.
These microbial mats carpet the bed of the lake, interspersed with sparse
stands of submerged macrophytes covered with periphyton in the shallow
areas and some deeper zones. Periphyton and the surface golden-brown layer
of the mats mainly comprised the same diatom species. This distribution gave
the appearance of a continuous vegetation dominated by periphytic diatoms
in contrast to the situation in 1989, when areas of microbial mats could be
clearly distinguished from stands of submerged macrophytes.
Fauna
Despite the relatively large number of publications concerning Chiprana
(Alonso, 1980, 1985a, 1985b; Martino, 1988; Alonso, 1990; Guerrero et al.,
1991; Berga et al., 1991; Vidondo and Guerrero, 1992; Amat et al., 1995.),
there is still a lack of information on major aspects at different temporal
scales. Special interest shown in the biology of Artemia parthenogenetica
has meant that its presence has been the best documented during previous
203
years. The species was determined incorrectly as Artemia salina (L.) by
Alonso (1980, 1985) and Martino (1988). The composition of zooplankton
communities has been assessed only once (Martino, 1988) prior to this study.
Two main communities can be distinguished, one associated with the
freshwater area close to the inflow of the drainage canal, bringing waste irrigation water (Herpetocypris chevreuxi (Ostracoda), Sigara stagnalis Leach
(Heteroptera) and Coelambus pallidulus (Coleoptera), and another one which
is the most distant from the influence of the canal, and composed of halobionts (Fabrea salina (Ciliata), Hexarthra fennica Schmarda and Brachionus
plicatilis (O.F. Müller) (Rotifera), Arctodiaptomus salinus (Daday) and
Cletocamptus retrogressus Schmankewitsch (Copepoda); Artemia parthenogenetica (Anostraca) and Diptera of the genus Ephydra).
Except for Artemia and a few cyclopoid copepods, the zooplankton and
zoobenthos communities recorded in 1994/95 comprised the same species as
those found previously by Martino (1988).
Artemia parthenogenetica appeared in May 1995, after several years of
absence from the lake (Alonso, pers. comm.), and was distributed relatively
homogeneously throughout the water-body away from the area influenced by
the freshwater inlet.
Another common member of the zooplankton and found mainly in the
saline area throughout the hydrological cycle was the calanoid Arctodiaptomus salinus. This underwent a massive bloom in July 1994, but unlike
Artemia parthenogenetica, it had always been recorded in all previous studies
of the lake.
The planktonic ciliate Fabrea salina was recorded as an abundant species
in both the littoral and deep water areas, and across the salinity gradient.
Other taxa were also recorded (acarids, nematodes and turbellarians),
but not identified to species. Finally, numerous ephippia (resting eggs) of
Daphnia were found in samples from the lake, but few adults were seen.
Discussion
The present state of the Salada de Chiprana indicates that a significant change
in its trophic pattern has taken place over the last 5 years. This is shown by the
increase in total-P and total-N concentration (Figure 3; Table 2). The causes
of these changes are presumably anthropogenic factors, namely an increased
inflow of freshwater rich in nutrients, which affected the biological equilibrium of the lake, in general, and the communities of primary producers, in
particular. Also, hydrological features of the periods involved could have had
an effect on the nutrient concentrations in the lake. In relation to this, a major
204
evaporative loss of water in 1994/95 compared to 1989 could have stimulated
the concentration of nutients during this period.
The marked increase in the mean concentrations of nutrients in the lake
during 1994/95 is believed to be one of the main factors explaining changes
in relation to primary producers. Fundamentally, P in the water column is
found mostly in the form of tightly bound chemical complexes, unavailable
to phytoplankton communities (Figure 4). Periphyton, however, showed an
explosive growth which could indicate conditions of phosphorus sufficiency
in the benthic zone. Periphyton, on the other hand can mobilize nutrients from
either the water column or sediments (Sand-Jensen et al., 1991) and therefore
possesses a competitive advantage. However, some laboratory experiments
suggest that the saturation of cellular uptake of phosphate in the periphyton
is affected by the thickness of its matrix. Thus, a concentration ten times
higher than 1 µg L−1 of PO4 (theorical saturation point) can be required
to maintain its growth rate (Bothwell, 1989). In relation to P levels in the
benthic zone, Caraco et al. (1989) proposed a model for systems with high
sulphate concentrations in which phosphate is released in the first millimeters
of sediments due to chemical competition with sulphate.
As the Salada de Chiprana is relatively shallow (Z max = 5m), light
penetrates to the bottom of the lake, allowing the development of benthic
autotrophic communities in conditions where both light and nutrients are
available.
The phytoplanktonic community may have decreased partly due to phosphorus limitation, but also from intense grazing pressure by zooplankton
(Jellison and Melack, 1988). Evidence of this occurred in August 1995 when
the water of the lake became transparent, an event coinciding with a bloom
of Artemia. Artemia can form dense populations with considerable powers of
filtration (Hammer, 1986; Jellison and Melack, 1988; Javor, 1989).
This dynamic phenomenon in the Salada may have caused important modifications to the physico-chemical environment of the water column. One of
the principal changes would be that the photosynthetic oxygen produced by
periphyton in the deepest areas of the lake would have substantially transformed the previously anaerobic conditions with high levels of H 2 S in bottom
waters of these areas. In the presence of oxygen, hydrogen sulphide liberated
from the sediment would change into oxidised forms. This great change in
the vertical structure of the water column would, in turn, reduce the electron
donor source for the process of photosynthesis and cause the displacement
of anaerobic phototrophic bacteria, which had principally been responsible
for the fixation of inorganic carbon in 1989. Another factor associated with
the disappearance of the anaerobic phototrophic bacteria is a reduction in the
205
shade effect which these created; subsequently this would have allowed the
growth of periphyton in the deepest areas.
Both the increase in nutrients and periphyton seem to have affected the
basic structure of the microbial mat, impeding the growth of some of the most
characteristic species and/or altering their relative position in the mat. The
disappearance of the cyanobacterium Microcoleus chthonoplastes (Vidondo,
1991) may indicate the increased loading of organic material as this species
is characteristic of oligotrophic environments (Golubíc, 1973). Furthermore,
in relation to the degree of compaction of sediments, Microcoleus has a
growth pattern which interlaces the different components of the microbial
mat (see Figure 7), and so the disappearance of this cyanobacterium may be
the fundamental cause of the loss of cohesion and compactness of the mat.
In appearance, the mats seem to be in a process of degradation: the increased
number of heterotrophic bacteria throughout the mat may indicate that the
dominant process is that of mineralisation, and this may also help to explain
the disappearance of Microcoleus chthonoplastes. The latter species has a
high rate of CO2 fixation (Guerrero and De Wit, 1992), and this not only
maintains the physical structure of the mat but also its metabolic processes.
Other alterations could equally affect the establishment of a different type
of microbial mat. These include the change in granulometric characteristics
of the sediment due to the increase of organic material which disfavours the
growth of complex, compact mats such as those dominated by, for example,
Microcoleus chthonoplastes (Florín, 1994; Guerrero et al., 1994).
Note also the disappearance of the red sulphur bacteria, Chromatium. The
increase in salinity during 1994/95 may explain the disappearance of this
taxon as Chromatium is found only in environments of moderate salinity and
is not a strong competitor with other genera of sulphur bacteria in hypersaline
conditions (Javor, 1989).
The poor development of macrophytes and charophytes is undoubtedly
related to the fact that they are completely covered with periphyton and
therefore unable to capture sufficient light for photosynthesis. As in the case
of microbial mats, another important factor is the increase in organic matter
and subsequent changes in sediment texture; this results in a loose, unstable
substrate where it is difficult for plants to root.
In relation to the animal community, the changes at the primary producer
level have not caused a significant change in either the structural or functional
diversity of animals.
The increased flow of drainage water towards the Salada, however, seems
to have altered existing biocenoces. On the one hand, it allowed colonisation
by more generalist and less halophilous taxa, and on the other, a change in the
206
phenology of the more halophilous species, such as Artemia which is believed
to have been absent since 1989.
The distribution of zooplankton and meiobenthos species clearly shows
the response of taxa to the inflow of fresh water. Artemia parthenogenetica, Arctodiaptomus salinus and Fabrea salina are all cosmopolitan species
typical of saline environments, and were all found in areas of the lake least
affected by the canal. Other species appear to be good indicators of the transition zone between fresh and saline water. Among these is Herpetocypris
chevreuxi, an oligohaline ostracod species found mainly in littoral zones
where salinity ranges from freshwater to slightly saline. Other species found
in the area affected by the canal are littoral species with a wide range of
salinity tolerance. It is difficult to determine whether their presence in this
area is related to salinity per se.
Artemia parthenogenetica is often used in laboratory experiments in cultures of different salinities, with maximum rates of survival at high salt
concentrations (75 g L−1 ) (Basil and Pandian, 1991). The parthenogenetic
population inhabiting Chiprana is tetraploid and adapted to extreme temperatures. Many polyploid animal species appear more tolerant of the stress of low
temperatures and possess a high capacity for colonization (Amat et al., 1995).
In the Iberian Peninsula, tetraploid parthenogenetic populations of Artemia
are distributed at latitudes above 40◦ N, whereas diploid parthenogenetic ones
are found below this latitude. The bisexual populations are isolated from
parthenogenetic ones and are found in both natural and artificial coastal
systems.
It is suggested that the disappearance of Artemia parthenogenetica sometime between 1989 and 1995 may have been due to physico-chemical changes
induced by the entry of fresh water to the lake. However, studies by Amat
in this lake in 1985 showed that Artemia can tolerate salinities of 40 gL−1 ,
approximately the lower limit of tolerance known for the species. Therefore
the role of salinity in the regulation of Artemia populations is not at all clear
in the present study, and the effects that pesticides and changes in trophic
state can have on the development of the species are yet to be resolved.
The type of response exhibited by the Salada de Chiprana to the change
in trophic pattern does not seem to agree with the equilibrium model of
turbid waters in shallow freshwater systems (Scheffer et al., 1993). Unlike
this model, the Salada responded to increased nutrient loading with a significant decrease in phytoplankton biomass. Similar phenomena are cited by
other authors in saline lakes (Robert et al., 1995). At the same time, the
displacement of submerged macrophytes and the dominance of periphyton
was apparently caused by other factors related to both the capacity to capture
light and to sediment characteristics. The dense growth of periphyton in this
207
Figure 8. Biotic and abiotic interactions between different components in the Salada de
Chiprana. Black arrows represent increasing relations and grey arrows, decreasing relations.
Main factors affecting the trophic state are represented by grey boxes.
situation would favour stability of the lake bed and restrict the resuspension of
sediments. As a result, an increase in turbidity of the water would be avoided,
and a positive feedback would develop whereby the extensive growth of
periphyton would tend to clarify the water by stabilisation of sediments and
favouring further growth through improved light conditions. Figure 8 represents the main interactions between biotic and abiotic components in the
Salada de Chiprana for 1994/95. To a large extent the “clear water” condition could also be from the effects of filtration by the zooplankton during a
large part of the hydrological cycle. These would liberate suspended particles
into water column but restrict the processes caused by eutrophication in the
benthic zones.
One of the most noticeable changes in the lake and one of concern was
the marked decrease in both absolute and relative diversity in the primary
producer community. Contrary to the generally assumed explanation for this
decline, the cause could not be related to the increase in salinity. It is related
more to the internal reorganisation of the phototrophic population in response
to perturbations caused by the increased nutrient loading, the effects of zooplankton grazing, and local climatic variables since the stability of the water
column is strongly dependent on wind and precipitation.
With this in mind, it appears that a new conceptual model of response to
perturbation is needed for hypersaline waters, where the capacity of resistance
208
and restoration of the ecosystem is modified by important chemical factors
and is comparatively greater than observed in freshwater systems.
The results of this study emphasize that the alterations in the Salada de
Chiprana seem to result more from modifications in nutrient load to the
lake, exacerbated by an extensive drought, than from dilution of salts by the
inflow of fresh water. This situation challenges the general idea that change
in salinity is the dominant factor causing alteration in permanent, hypersaline ecosystems. It appears that, at least in Salada de Chiprana, changes in
trophic pattern appear to have a much greater effect on the primary producers
community than changes in salinity.
These results are important in the establishment of new criteria for the
evaluation of and appropriate conservation measures applicable to inland
saline systems, as changes which are generally assumed to be caused entirely
by the dilution of saline systems may in fact be masking more important ones
caused by the simultaneous input of nutrients.
Acknowledgements
Special thanks to Mr Geoff Oliver and Mrs Irma Vila for the English revision . The study was funded by the Spanish Interministerial Commission of
Science and Technology (ref. AMB94-0827). Paula Díaz was supported by a
scholarship of the Instituto de Cooperación Iberoamericana (ICI).
References
APHA, AWWA and WPCF 1992. Métodos normalizados para el análisis de las aguas potables
residuales (Standard Methods for the examination of water and waste water). Ediciones
Díaz de Santos, S. A.
Amat, F., Barata, C. and Hontoria, I. 1995. Biogeography of genus “Artemia” (Crustacea,
Branchiopoda, Anostraca) in Spain. International Journal of Salt Lake Research 3: 175–
190.
Alonso, M. 1980. Nota sobre algunes Ctenodaphnidae de la Península Ibérica. Bullt. Inst. Cat.
Nat. 45: 470–53.
Alonso, M. 1985a. Las lagunas de España Peninsular. Taxonomía, ecología y distribución de
los cladoceros. Ph. D. Thesis. Universidad de Barcelona, 795 pp.
Alonso, M. 1985b. A survey of the Spanish euphyllopoda. Misc. Zool. 9: 179–208.
Alonso, M. 1990. Anostraca, Cladocera and Copepoda of Spanish lakes. Hydrobiologia 197:
221–231.
Aminot, A. and Chaussepied, M. 1983. Manuel des analyses chimiques en milieu marin.
CNEX, Paris.
Barko, J.W., Gunnison, D. and Carpenter, S.R. 1991. Sediment interactions with submersed
macrophyte growth and community dynamics. Aquatic. Botany 41: 41–65.
209
Alonso, M. 1987. Clasificación de los complejos palustres españoles. Bases Científicas para
la Protección de los Humedales en España, Real Academia de Ciencias Exactas Físicas y
Naturales.
Basil, A. and Pandian, G.T. 1991. Culturing Artemia (Tuticorin strain) in organic and
agricultural wastes at different salinities. Hydrobiologia 212: 11–17.
Berga, A., Cabrera, M. and Legaz, A. 1994. Medidas correctoras en el funcionamiento hídrico
de la Laguna Salada de Chiprana. Actas de las “Jornadas sobre el futuro de las Saladas de
los Monegros y el Bajo Aragón”. Ediciones del Grupo Cultural Caspolino.
Bothwell, M.L. 1989. Phosphorus-limited growth dynamics of lotic periphyton diatom communities: areal biomass and cellular growth rate responses. Canadian Journal Fisheries
and Aquatic Science 46: 1293–301.
Campbell, C. and Prepas, E. 1986. Evaluation of factors related to the unusually low chlorophyll levels in prairie saline lakes. Canadian Journal Fisheries and Aquatic Science 43:
846–854.
Caraco, N., Cole, J. and Likens, G.E. 1989. Evidence of sulphate-controlled phosphorus
release from sediments of aquatic systems. Nature (Lond.) 341: 316–318.
Florín, M. 1994. Funcionamiento de las lagunas salinas temporales manchegas. Relación entre
las fluctuaciones hídricas, hidroquímicas y dinámica trófica. Tésis doctoral, Universidad
Autónoma de Madrid.
Friend, P., Hirst, J. and Nichols, G.J. 1986. Sandstone body structure and river process in the
Ebro basin Aragón, Spain. Cuadernos Geología Ibérica 10: 9–30.
Geitler, L. 1932. Kryptogamen flora von Deutschland, Österriech und der Schweiz:
Cyanophyceae von Europa. Akademishe Verlangsgesellschaft m.b.H., Leipzig.
Guerrero, M.C., Balsa, J., Pascual, M.L., Martínez, B. and Montes, C. 1991. Caracterización
limnológica de la laguna salada de Chiprana (Zaragoza, España) y sus comunidades de
bacterias fototróficas. Limnetica 7: 83–96.
Guerrero, M.C. and De Wit, R. 1992. Microbial mats in the inland saline lakes of Spain.
Limnetica 8: 197–204.
Guerrero, M.C., Tadeo, A.B. and de Wit, R. 1994. Environmental factors controlling the development of microbial mats in inland saline waters. The granulometric composition of the
sediment. In: L.J. Stal and P. Caumette (Eds) Microbial Mats, pp. 85–90. Springer-Verlag,
Berlín Heidelberg.
Gulobíc, S. 1973. The relationship between blue-green algae and carbonate deposits. In: N.G.
Carr and B.A. Whitton (Eds) The Biology of Bluegreen Algae, pp. 434–472. Blackwell
Scientific Publications Oxford.
Hammer, T. 1984. The saline lakes of Canada. In: F.B. Taub (Ed) Lakes and Reservoirs,
pp. 521–539. Elsevier Science Publishers B. V., Amsterdam.
Hammer, T. 1986. Saline lakes of Canadian prairies. Revue Canadienne des Ressources en
Eau 11 (1): 43–57.
Javor, B. 1989. Hypersaline Environments: Microbiology and Biogeochemistry. SpringerVerlag, Berlin, 328 pp.
Jeffrey, S. and Humphrey, G.F. 1975. New spectrophotometric equations for determining
chlorophylls a y b and c and c1 in high plants, algae and natural phytoplankton. Biochemie
Physiologie Pflanzen 167: 1991–1994.
Jellison, R. and Melak, J.M. 1988. Photosynthetic activity of phytoplankton and its relation to
environmental factors in hypersaline Mono Lake, California. Hydrobiologia 158: 69–88.
Jorgensen, B., Revsbech, N. and Cohen, Y. 1983. Photosynthesis and structure of benthic
microbial mats. Microelectrode and SEM studies of four cyanobacterial communities.
Limnology and Oceanography 28: 1075–1093.
210
Menéndez, M., Forés, E. and Comín, F.A. 1993. Decomposition of Ruppia cirrhosa (Petagna)
Grande in the sediment of a coastal lagoon. Sci. Mar. 57(1): 15–21.
Martino, P., Montes, C. and Alonso, M. 1986 a. Las lagunas saladas de España: Una fuente de
recursos poco conocida. Actas de las Jornadas sobre la conservación de la naturaleza en
España: Naturaleza y Sociedad, Caja de Ahorros de Asturias (Eds), pp. 231–234.
Martino, P., Montes, C. and Alonso, M. 1986b. Sobre la conservación y gestión de las lagunas
saladas de España. Actas de las Jornadas sobre la conservación de la naturaleza en España:
Naturaleza y Sociedad, Caja de Ahorros de Asturias (Eds), pp. 235–238.
Martino, P. 1988. Limnología de las lagunas salinas españolas. Tésis doctoral, Universidad
Autónoma de Madrid.
Montes, C. and Martino, P. 1987. Las lagunas salinas españolas. En: Bases científicas para
la protección de humedales en España. Real Academia de Ciencias Exactas Físicas y
Naturales, Madrid.
Plummer, J., Jones, B. and Truesdell, A. 1984. WATEQ-A FORTRAN IV version of WATEQ,
a computer program for calculating chemical equilibrium of natural waters. Water
Resources Investigations U.S. Geological Survey 76: 31.
Por, F. 1980. A classification of hypersaline waters based on trophic criteria. Mar. Biol. 1:
121–131.
Riba, O., Villena, J. and Quirantes, J. 1967. Nota preliminar sobre la sedimentación en paleocanales terciarios de la zona de Caspe-Chiprana (Zaragoza). Anal. Edaf. Agrob. 26:
617–634.
Robarts, R.D., Evans, M.S. and Arts, M. 1992. Light, nutrients and water temperature as
determinants of phytoplankton production in two saline, prairie lakes with high sulphate
concentrations. Canadian Journal Fisheries and Aquatic Science 49: 2281–2290.
Robarts. R.D., Doland, D. and Arts, M. 1995. Phytoplankton primary production of three
temporary prairie wetlands. Canadian Journal Fisheries and Aquatic Science 52: 897–902.
Rippka, R. Deruelles, J. Waterburry, J. Herdman, M. and Stanier, R. 1979. Genetic assignements, strain histories and properties of the pure cultures of cyanobacteria. Journal of
General Microbiology 111: 1–61.
Sand-Jensend, K. and Borum, J. 1991. Interactions among phytoplankton, periphyton, and
macrophytes in temperate freshwaters and estuaries. Aquatic Botany 41: 137–175.
Sheffer, M., Hosper, S., Meijer, M., Moss and Jeppensen, E. 1993. Alternative equilibria in
shallow lakes. TREE 8: 275–279.
Vidondo, B. 1991. La laguna Salada de Chiprana: Descripción de sus características fiicoquímicas como hábitat para sus singulares comunidades de bacterias fototrofas. Memoria
de licenciatura. Universidad Autónoma de Madrid.
Vidondo, B. and Guerrero, M.C. 1992. La laguna Salada de Chiprana: Descripción de sus
características fisicoquímicas como hábitat para sus singulares comunidades de bacterias
fototrofas. Cuaderno de estudios Caspolinos XVIII, 117–246.
Vidondo, B., Martínez, B., Montes, C. and Guerrero, M.C. 1993. Physico-chemical characterization of a permanent Spanish hypersaline lake: La Salada de Chiprana. Hydrobiologia
267: 113–125.
Walter, M.R. (Ed) 1976. Stromatolites.Developments in Sedimentology. 20, Elsevier, Amsterdam.
Williams, W.D. 1981. Problems in the management of inland saline lakes. Verh. Internat.
Verein. Limnol. 21: 688–692.
Williams, W.D. 1993. Conservation of salt lakes. Hydrobiologia 267: 291–306.