Major changes in CO2 efflux when shallow lakes shift from a turbid to a clear water state

Abstract

Lakes can be sources or sinks of carbon, depending on local conditions. Recent studies have shown that the CO 2 efflux increases when lakes recover from eutrophication, mainly as a result of a reduction in phytoplankton biomass, leading to less uptake of CO 2 by producers. We hypothesised that lake restoration by removal of coarse fish (biomanipulation) or invasion of mussels would have a similar effect. We studied 14–22 year time series of five temperate Danish lakes and found profound effects on the calculated CO 2 efflux of major shifts in ecosystem structure. In two lakes, where limited colonisation of submerged macrophytes occurred after biomanipulation or invasion of zebra mussels (Dreissena polymorpha), the efflux increased significantly with decreasing phytoplankton chlorophyll a. In three lakes with major interannual variation in macrophyte abundance, the efflux declined with increasing macrophyte abundance in two of the lakes, while no relation to macrophytes or chlorophyll a was found in the third lake, likely due to high groundwater input to this lake. We conclude that clearing water through invasive mussels or lake restoration by bioma-nipulation may increase the CO 2 efflux from lakes. However, if submerged macrophytes establish and form dense beds, the CO 2 efflux may decline again.

Full text

SHALLOW LAKES
Major changes in CO
2
efflux when shallow lakes shift
from a turbid to a clear water state
Erik Jeppesen .Dennis Trolle .Thomas A. Davidson .Rikke Bjerring .
Martin Søndergaard .Liselotte S. Johansson .Torben L. Lauridsen .
Anders Nielsen .Søren E. Larsen .Mariana Meerhoff
Received: 14 April 2015 / Revised: 26 August 2015 / Accepted: 29 August 2015 / Published online: 23 October 2015
ÓSpringer International Publishing Switzerland 2015
Abstract Lakes can be sources or sinks of carbon,
depending on local conditions. Recent studies have
shown that the CO
2
efflux increases when lakes recover
from eutrophication, mainly as a result of a reduction in
phytoplankton biomass,leading to less uptake of CO
2
by
producers. We hypothesised that lake restoration by
removal of coarse fish (biomanipulation) or invasion of
mussels would have a similar effect. We studied
14–22 year time series of five temperate Danish lakes
and found profound effects on the calculated CO
2
efflux
of major shifts in ecosystem structure. In two lakes,
where limited colonisation of submerged macrophytes
occurred after biomanipulation or invasion of zebra
mussels (Dreissena polymorpha), the efflux increased
significantly with decreasing phytoplankton chlorophyll
a. In three lakes with major interannual variation in
macrophyte abundance, the efflux declined with
increasing macrophyte abundance in two of the lakes,
while no relation to macrophytes or chlorophyll awas
found in the third lake, likely due to high groundwater
input to this lake. We conclude that clearing water
through invasive mussels or lake restoration by bioma-
nipulation may increase the CO
2
efflux from lakes.
However, if submerged macrophytes establish and form
dense beds, the CO
2
efflux may decline again.
Keywords Air–water CO
2
flux Recovery 
Eutrophication Macrophytes Zebra mussel Lake
metabolism
Guest editors: M. Bekliog
˘lu, M. Meerhoff, T. A. Davidson,
K. A. Ger, K. E. Havens & B. Moss / Shallow Lakes in a Fast
Changing World
Electronic supplementary material The online version of
this article (doi:10.1007/s10750-015-2469-9) contains supple-
mentary material, which is available to authorized users.
E. Jeppesen (&)D. Trolle T. A. Davidson 
R. Bjerring M. Søndergaard L. S. Johansson 
T. L. Lauridsen A. Nielsen M. Meerhoff
Lake Ecology Section, Department of Bioscience and the
Arctic Research Centre, Aarhus University, Aarhus,
Denmark
e-mail: ej@dmu.dk; ej@bios.au.dk
E. Jeppesen D. Trolle T. L. Lauridsen S. E. Larsen
Sino-Danish Centre for Education and Research (SDC),
Beijing, China
S. E. Larsen
Catchment Science and Environmental Management
Section, Department of Bioscience, Aarhus University,
Aarhus, Denmark
S. E. Larsen M. Meerhoff
Departamento de Ecologı
´a Teo
´rica y Aplicada, Centro
Universitario de la Regio
´n Este-Facultad de Ciencias,
Universidad de la Repu
´blica, Maldonado, Uruguay
123
Hydrobiologia (2016) 778:33–44
DOI 10.1007/s10750-015-2469-9

Introduction
Lakes can act as either sources or sinks of carbon
depending, among several processes, on the balance
between photosynthesis and ecosystem respiration as
well as the external nutrient input (Cole et al., 1994,
2000). The importance of lakes in the global carbon
cycle is not yet fully resolved (Moss et al., 2011;
Raymond et al., 2013) because most models have
typically focused on marine, terrestrial and atmo-
spheric compartments. However, our understanding of
the role of freshwaters as a source or sink of CO
2
is
increasing (Dean & Gorham, 1998; Stallard, 1998;
Cole et al., 2007, Cole, 2013; Barros et al., 2011).
It is now evident that inland waters, and lakes in
particular, can play an important role in CO
2
exchange
with the atmosphere (Tranvik et al., 2009). Lakes are
often supersaturated with CO
2
(e.g. Kling et al., 1992;
Cole et al., 1994; Kosten et al., 2010), thereby acting as
aCO
2
source to the atmosphere. The saturation level is
greatly influenced by primary production, which
depletes CO
2
through photosynthesis and increases
pH (Talling, 1976). With increasing eutrophication,
phytoplankton productivity increases and water CO
2
decreases (Wetzel, 2001). Furthermore, more carbon
is buried in the lake sediments in eutrophic lakes
(Anderson et al., 2014) and the CO
2
efflux is therefore
reduced. Conversely, the efflux is expected to increase
again as lakes recover from eutrophication (Provoost
et al., 2010) after reductions in the external nutrient
input. For instance, substantial reductions in the
nutrient loading to four Danish lakes resulted in an
increase in annual average CO
2
effluxes of[30% over
a 20-year period, concurrently with decreasing phyto-
plankton biomass (Trolle et al., 2012). Moreover, the
largest changes in the CO
2
efflux were observed in
winter when the reduction in chlorophyll a(Chl a) was
most pronounced, while the summer efflux was lower
and at times negative as internal phosphorus loading
kept phytoplankton biomass relatively high during the
warmer months (Trolle et al., 2012).
Shifts in trophic structure may ultimately affect the
composition and nature of primary producers and
consequently net productivity, water pH and CO
2
efflux.
In an attempt to obtain clear water conditions following
nutrient loading reduction, plankti-benthivorous fish may
be removed (biomanipulation) to enhance zooplankton
grazing on phytoplankton and to reduce physical
disturbance of the sediments (Carpenter et al., 1998;
Jeppesen et al., 2012), thereby speeding up recovery. A
reduction in phytoplankton Chl awould be expected to
result in higher CO
2
efflux, but, at least in the short term,
one might also expect that a reduction in disturbance of
the sediment would reduce decomposition and enhance
net accumulation in the sediment, potentially counter-
acting the increased CO
2
efflux.
Phytoplankton biomass and production may also be
notably affected by other changes in the trophic web,
such as the introduction of invasive key-stone species.
Over the last few decades, many lakes worldwide
have been colonised by invasive bivalves, including
golden mussel (Limnoperna fortune (Dunker, 1857))
and zebra mussel (Dreissena polymorpha (Pallas,
1771)) (e.g. Strayer et al., 1999; Karatayev et al.,
2007). The establishment of zebra mussels, in partic-
ular, is often followed by major changes, particularly
reduced phytoplankton biomass and increased water
clarity (Idrisi et al., 2001; Mayer et al., 2002; Higgins
& Vander Zanden, 2010). As with biomanipulation,
such changes can lead to higher CO
2
efflux although
faeces from the mussels might enhance net carbon
accumulation (Stewart et al., 1998).
In turn, improving water clarity in shallow lakes
provides an opportunity for submerged macrophytes
to increase in abundance or (re)colonise (Moss, 1990;
Moss et al., 1996; Scheffer et al., 1993). In dense
macrophyte stands, water circulation can be low and
the concentration of CO
2
may drop to very low levels
during daytime, most markedly in the upper part of the
macrophyte stand (Jones et al., 1996). Many sub-
merged macrophytes can use bicarbonate (HCO
3), but
CO
2
is generally the preferred carbon source (Jones,
2005). CO
2
uptake by the macrophytes and associated
periphyton can also lead to precipitation of calcium
carbonate (CaCO
3
). Moreover, lakes with abundant
macrophytes may store additional carbon due to
accumulation of plant material in the sediments
(Carpenter, 1981). Therefore, presence of submerged
plants may increase the likelihood of a lake being a
carbon sink. We therefore hypothesised that although
the CO
2
efflux should increase in eutrophic lakes in the
early phase after nutrient loading reduction, owing to
reduction in phytoplankton biomass, this process
should be reversed if submerged macrophytes become
abundant and permanently established. Very high
cover and biomasses of submerged macrophytes under
34 Hydrobiologia (2016) 778:33–44
123

mesotrophic to eutrophic conditions are only possible
in shallow lakes (Wetzel, 2001) where the water depth
is potentially low enough to allow light penetration to
the bottom over large areas. Thus, it may be hypoth-
esised that the positive effect on carbon sequestration
of macrophyte establishment following improvements
in water clarity is mainly of importance in shallow
lakes.
To investigate the effect of substantial changes in
trophic structure on the CO
2
efflux from temperate
shallow lakes, we estimated the CO
2
efflux using data
from long-term studies of five Danish lakes before and
after major shifts in trophic structure (fish removal,
macrophyte biomass changes and zebra mussel
colonisation).
Methods
Study sites
All five study lakes are eutrophic, shallow and
polymictic. Basic data are given in Table 1. Lake
Faarup was colonised by zebra mussels in the 1990s,
mussel larvae being observed for the first time in 1993,
and veliger larvae have been present in the plankton
from 1998 onwards (Jeppesen et al., 2012). In 2000,
their density was up to 1,300 ind. m
-2
. Since 1995, a
major decrease has occurred in summer mean Chl a,
TP, TN, and annual mean Chl a, but not in annual
mean TP and TN (Jeppesen et al., 2012). Accordingly,
water transparency measured as Secchi depth has
increased, but macrophytes have very low cover. As
the external loading of TN and TP has not changed
during the study period, the drastic changes can be
attributed to the colonisation and a gradual increase in
zebra mussel densities (Jeppesen et al., 2012).
Lake Væng receives about 90% of its water from
groundwater (Kidmose et al., 2013) and the water
residence time is very short (Table 1). For many years,
the lake received waste water from a local community.
Following diversion of this loading in 1981, the lake
remained turbid (Søndergaard et al., 1990) and was
therefore biomanipulated during 1986–1988 by
removing approximately 50% (4 tonnes) of the fish
stock (almost exclusively bream, Abramis brama L.,
and roach, Rutilus rutilus L.). Subsequently, phyto-
plankton Chl aand water turbidity decreased substan-
tially and the lake was colonised by submerged
macrophytes, first Potamogeton crispus L. and then
Elodea canadensis Michx., the latter completely
covering the lake within a 1–2 year period (Lauridsen
et al., 1994). However, from 1998 onwards, macro-
phytes largely disappeared. Concurrently, the abun-
dance of roach and small perch (Perca fluviatilis L.)
increased (Jeppesen et al., 2012). In an attempt to shift
the lake back to the clear state, 2.7 tonnes (68% of the
first biomanipulation effort) of bream and roach were
removed during 2007–2009. The concentration of Chl
adecreased and the coverage of submerged macro-
phytes increased from \1% to ca. [70%, and since
2009 the biomasses of roach and bream have remained
low compared with the years prior to the fish
manipulation (Jeppesen et al., 2012).
Lake Engelsholm was also subjected to nutrient
loading reduction and showed delayed recovery
(Bjerring et al., 2013). Restoration by biomanipulation
was conducted in 1992–1994. Nineteen tonnes of
cyprinid fishes were removed and their estimated
biomass subsequently decreased from 675 to
150–300 kg ha
-1
(Jeppesen et al., 2012). Biomanip-
ulation led to a substantial reduction of phytoplankton
Chl a, total phosphorus (TP) and total nitrogen (TN),
as well as an increase in Secchi depth (Liboriussen
et al., 2007). Submerged macrophyte coverage has
Table 1 Physical data on the five study lakes including information on phenomena studied
Lake Lake
area
(km
2
)
Lake
catchment
area (km
2
)
Lake
mean
depth (m)
Lake
maximum
depth (m)
Water
residence
time (years)
Events
Faarup 0.99 13.2 5.6 11.1 0.6 Zebra mussel invasion
Væng 0.16 9 1.2 1.9 0.06 Biomanipulation
Engelsholm 0.44 15.2 2.4 6.1 0.73 Biomanipulation
Arreskov 3.17 24.8 1.9 3.6 1.1 Fish kill and biomanipulation
Stigsholm 0.21 0.8 1.2 0.015 Natural variation in plant coverage
Hydrobiologia (2016) 778:33–44 35
123

generally been very low, both before and during the
early years following biomanipulation. However,
during 2007–2010 the coverage increased gradually
from 2 to 12%.
Lake Arreskov received untreated sewage from the
1950s to the 1980s, leading to severe eutrophication.
Sewage was diverted in 1983. A major fish kill
occurred during autumn and winter 1991. Four tonnes
of cyprinids were removed in 1995, and the lake was
stocked with under-yearling piscivores (pike, Esox
lucius L.), in total 141 ind ha
-1
in 1993 and 1995.
Cyprinid biomass was calculated as 172 kg ha
-1
in
1987 and 71 kg ha
-1
in 1995 (Sandby & Hansen,
2007). In general, Lake Arreskov has shown large
inter-annual alternations in ecological state (sensu
Scheffer et al., 1993), exhibiting high water clarity and
high abundance of submerged vegetation in some
years, for instance 1996–1998 and 2003–2011, and
limited submerged vegetation and, consequently,
turbid water and dominance of cyanobacteria in other
years, for example 1995 and 1999–2002 (Nielsen
et al., 2014). The dominant submerged plant species
were Zannichellia palustris L., Potamogeton crispus
L., Potamogeton pectinatus L., Chara vulgaris L. and
Chara globularis L.. In other years, Ceratophyllum
demersum L. made a large contribution to the biomass.
Lastly, Lake Stigsholm has also fluctuated between
alternative states (i.e. a macrophyte-rich clear water
state and a macrophyte-poor turbid state, Scheffer
et al., 1993) since the 1980s, the dominant submerged
macrophyte species being P. pectinatus, P. berchtoldii
L. Callitriche sp. and Elodea canadensis L. as well as
filamentous algae (Jeppesen et al., 1992, 1998;
Søndergaard et al., 1998).
Chemical analyses
In all lakes depth-integrated samples were taken from
the photic zone, minimum 19 times annually, and
analysed according to standard methods (Kronvang
et al., 1993). Basic chemical data on the lakes are
given in Table 2.
Submerged macrophyte coverage
Depending on lake area, aquatic macrophytes were
inspected at 30–200 equidistant observation points
situated along transects covering the entire lake area.
At each observation point, percentage macrophyte
coverage and species composition were determined
using an underwater viewing box. Total macrophyte
coverage for the lake (percentage area covered, PAC)
was calculated based on the total number of equidis-
tant observations and a species list was compiled.
Observations were made once a year between 1 July
and 15 August.
Calculation of CO
2
flux across the air–water
interface
The flux of CO
2
across the air–water interface (JCO2),
which can be influx or efflux for under-saturated and
super-saturated waters, respectively, was estimated
based on the diffusion film model (Stumm & Morgan,
1996); a technique that has been documented in more
detail in Trolle et al. (2012) and also outlined in the
Supplementary Material:
JCO2ðmol cm2h1Þ¼kb
CO2ðaqÞ

KHPCO2ðairÞ

;
where kis a transport coefficient (cm h
-1
), bis a factor
expressing the chemical enhancement of diffusion,
[CO
2(aq)
] is the concentration of dissolved CO
2
(mol
l
-1
, equivalent to 10
-3
mol cm
-3
), K
H
is Henry’s law
constant (mol l
-1
atm
-1
) and PCO2ðairÞis the partial
pressure of CO
2
in the atmosphere (atm).
Following the approach of Trolle et al. (2012), the
transport coefficient (k) was estimated from relations
to wind speed (based on the average of three individual
empirical relations). The chemical enhancement of
diffusion (b), which occurs at high pH and at low wind
speeds when the stagnant boundary layer is thick, was
estimated from water temperature, pH, ionic strength
and wind speed using the approach of Bade and Cole
(2006). The actual CO
2
concentrations in water
samples were estimated from pH and alkalinity
following Kling et al. (1992) and Cole et al. (1994).
Monthly averages of atmospheric PCO2ðairÞduring
the period 1989–2010 were acquired from the Earth
System Research Laboratory (ESRL) at the Mauna
Loa Observatory, Hawaii (http://www.esrl.noaa.gov/
gmd/ccgg/trends/). Monthly averages of wind speed
during the period 1989–2010, based on a 20 920 km
data grid covering all of Denmark, were used to derive
wind speeds at the location of each individual lake.
These monthly averages were used to calculate CO
2
36 Hydrobiologia (2016) 778:33–44
123

Table 2 Average annual CO
2
efflux and environmental variables from the five lakes studied. Means of annual means are shown with SD in parenthesis
Lake Faarup Væng Engelsholm
Period Before colonisation
(1989–1992)
During early
colonisation
(1993–1995)
When
established
1996–2003)
Low
macrophyte
years
(PAC \5%)
High
macrophyte
years
(PAC [5%)
Before
biomanipulation
(1989–1991)
During
biomanipulation
(1992–1993)
After
biomanipulation
(1994–2010)
Number of years 3 3 8 9 10 3 2 17
CO
2
efflux (mg C m
-2
d
-1
) 213 (75) 247 (25) 576 (271) 716 (244) 1046 (379) 288 (219) 424 (91) 698 (295)
pH 8.37 (0.03) 8.33 (0.04) 8.15 (0.30) 7.75 (0.40) 7.89 (0.32) 8.32 (0.27) 8.29 (0.02) 8.02 (0.10)
Alkalinity (mmol l
-1
) 1.99 (0.07) 1.98 (0.07) 2.05 (0.06) 1.16 (0.07) 1.12 (0.10) 1.56 (0.15) 1.54 (0.01) 1.60 (0.08)
Chlorophyll a(ug l
-1
) 42 (7) 42 (11) 21 (10) 52 (8) 36 (25) 55 (29) 73 (6) 27 (6)
Total phosphorus (ug l
-1
) 96 (11) 90 (13) 74 (4) 96 (8) 83 (27) 98 (30) 110 (7) 53 (10)
Total nitrogen (mg l
-1
) 1.5 (0.1) 1.5 (0.1) 1.3 (0.2) 0.9 (0.2) 0.7 (0.3) 2.4 (0.7) 2.4 (0.2) 1.2 (0.2)
Submerged macrophyte
coverage, PAC (%)
\0.5 0.6 (0.3) 0.5 (0.4) 0.7(1.3) 52.5 (29.0) \1\1 1.0 (1.3)
Lake Arreskov Stigsholm
Period Before biomanipulation After biomanipuation
(PAC \5%)
After biomanipulation
(PAC [5%)
Low macrophyte years
(PAC \5%)
High macrophyte years
(PAC [5%)
Number of years 2 4 11 4 16
CO
2
efflux (mg C m
-2
d
-1
) 368 (56) 472 (124) 283 (192) 350 (155) 141 (135)
pH 8.59 (0.06) 8.32 (0.14) 8.48 (0.20) 8.28 (0.21) 8.49 (0.30)
Alkalinity (mmol l
-1
) 2.30 (0.20) 2.58 (0.09) 2.35 (0.28) 1.09 (0.09) 1.11 (0.06)
Chlorophyll a(ug l
-1
) 134 (12) 38 (19) 66 (38) 48 (10) 37 (19)
Total phosphorus (ug l
-1
) 230 (3) 111 (30) 119 (56) 98 (18) 89 (24)
Total nitrogen (mg l
-1
) 3.5 (0.6) 2.0 (0.3) 2.1 (0.6) 2.5 (0.2) 2.6 (0.3)
Submerged macrophyte coverage, PAC (%) \1 1.8 (1.8) 27.1 (19.7) 1.6 (1.1) 30.8 (23.9)
Hydrobiologia (2016) 778:33–44 37
123

fluxes on individual sampling dates, in the corre-
sponding year and month and location, after which
linear interpolation between sampling dates was used
to generate monthly averages of the CO
2
flux for the
individual lakes.
Statistical analysis
Based on monthly means, average data from May–
September (summer), October–April (winter) and
annually were analysed. We used Pearson correlation
and multiple linear regressions to analyse changes in
fluxes in relation to different environmental variables.
In order to include carry-over effects from 1 year to the
next in the regressions, the time series data were
modelled using an autoregressive process of order 1,
which means that the output of the process depends on
the previous term in the process and the error term.
The SAS procedures MODEL and AUTOREG were
applied to estimate parameters for a number of
different linear models. In the multiple regressions,
alkalinity and pH were left out as these variables were
used in the CO
2
efflux calculations.
Results
All lakes released CO
2
annually, and in the winter
season during all study years, most notably in the
largely groundwater-fed L. Væng (Fig. 1; Table 2). In
all lakes, the CO
2
efflux was significantly higher (ttest
with auto-regression included, AR(1), P\0.01) in the
colder season, from October to April, than in summer,
particularly at high pH (Fig. 1). On average, the CO
2
efflux per day was from 2.9 (L. Arreskov) to 4.4 (L.
Væng) times higher in the colder season than in
summer, and in most years in L. Stigsholm a shift
occurred from efflux in the cold season to influx of
CO
2
in summer (Fig. 1).
Multiple regressions (with auto-regression, AR(1)
included) relating the CO
2
efflux to the abundance of
the two types of primary producers: phytoplankton
biomass (expressed as Chl a) and summer macrophyte
coverage, as well as water temperature, were con-
ducted separately at summer, cold season and annual
scales (Table 3). CO
2
efflux was significantly nega-
tively related to Chl aat both summer and annual
scales in the two lakes where macrophytes did not
make a significant contribution to primary producer
biomass (i.e. L. Faarup subjected to zebra mussel
invasion and biomanipulated L. Engelsholm). When
macrophyte coverage was extensive, albeit variable,
there was a significant negative relationship between
cover and the CO
2
efflux (i.e. L. Stigsholm and L.
Arreskov). In contrast, no effect of macrophytes and a
marginal effect of phytoplankton Chl awere found in
biomanipulated L. Væng. For L. Arreskov, Chl aalso
contributed negatively to the summer CO
2
efflux.
Temperature was not retained in any of the models.
Outside the growing season (October to April) the
pattern was less clear, Chl abeing significant for L.
Faarup and PAC (summer values) for L. Væng and L.
Arreskov, respectively (Table 3).
Discussion
Reduction in fish abundance or zebra mussel coloni-
sation and associated increases in macrophyte abun-
dance had pronounced effects on the CO
2
efflux in the
lakes studied. In the two lakes without macrophyte
colonisation (L. Engelsholm, L. Faarup), we recorded
a major increase in the efflux of CO
2
associated with
the reduction in phytoplankton biomass (Chl a) and
thus in pH (Tables 2,3). High carbon release associ-
ated with either fish removal or zebra mussel coloni-
sation was maintained in years with low Chl a(Fig. 1),
indicating that a higher efflux was not simply a
temporary effect related to a sudden shift in ecosystem
structure.
A different picture was found for lakes where
macrophytes were or became abundant. In L.
Stigsholm, exhibiting large inter-annual variations in
macrophyte abundance, the CO
2
efflux was negatively
related to the coverage of submerged macrophytes
both in summer and annually. In L. Arreskov, the
release was also lower in years with high macrophyte
coverage. These are conservative estimates as the
macrophytes reduce turbulence, likely leading to
progressive overestimation of the flux with increasing
macrophyte coverage; thus, the actual effect of the
plants might be stronger. High macrophyte and
periphyton production may lead to CO
2
depletion in
the water (Jones et al., 2000). Such a CO
2
depletion
and the plant-induced reduced turbulence could result
in CO
2
influx to the lake. Supporting this view, Boll
et al. (2012) found marked inter-annual variation in
the d
13
C (indicative of the sources of carbon to the
38 Hydrobiologia (2016) 778:33–44
123

ecosystem) of zooplankton and benthic macroinver-
tebrates in L. Væng. d
13
C was higher in years with
high abundance of macrophytes than when phyto-
plankton dominated the systems. They argued that this
reflected higher CO
2
consumption by plants and
associated periphyton in macrophyte-dominated
years, resulting in low CO
2
concentrations in the
water, which in turn would lead to less discrimination
against
13
C in photosynthesis (Peterson & Fry, 1987)
and, accordingly, to a higher d
13
C content of all
primary producers, including phytoplankton and then
consumers. However, use of bicarbonate (more
enriched in
13
C than atmospheric CO
2
) by the
macrophytes may also increase d
13
C. Others have
also observed higher d
13
C of primary producers when
macrophytes are abundant (Gao et al., 2014;de
Kluijver et al., 2015). In L. Væng the relationships
between the CO
2
efflux and Chl aand PAC were weak
despite major variations in PAC during the study
period, and only outside the growing season did the
high macrophyte coverage reduce the CO
2
efflux
(Fig. 1; Table 3). Mass mortality of macrophytes
during several winters followed by a major increase
in phytoplankton (Søndergaard et al., 1997; Lauridsen,
unpublished data) may have contributed to a negative
effect of macrophytes (measured during summer) on
the CO
2
efflux during this season. The relatively weak
response in this lake may reflect a very high input of
Fig. 1 Inter-annual variation in CO
2
efflux during summer
(May–September—dark blue dots) and outside the summer
season (October to April—light blue circles) in five lakes
subjected to: zebra mussel (D. polymorpha) invasion (L.
Faarup), restoration by biomanipulation (L. Væng, L. Engel-
sholm and L. Arreskov) and natural large variations in
macrophyte coverage (L. Stigsholm). Also shown are various
environmental variables (pH, chlorophyll a) and submerged
macrophyte coverage in late summer (July–Sept) as % of lake
area. The vertical grey areas show the timing of the major
changes leading to a shift in the lake ecosystems (D. polymorpha
veliger larvae present in the plankton in L Faarup; fish removal
in L. Væng and L. Engelsholm; fish kills in L. Arreskov)
Hydrobiologia (2016) 778:33–44 39
123

CO
2
-containing groundwater (Kidmose et al., 2013),
enhancing the CO
2
efflux (Fig. 1; Table 1) (Jeppesen
et al., 2012).
The effect of increasing water clarity, through
biomanipulation or zebra mussel invasion, on the
CO
2
balance was evident, but may differ between
shallow and deep lakes (Fig. 2). In shallow lakes, a
shift from a turbid to a clear water state may enhance
macrophyte and associated periphyton growth (Sch-
effer et al., 1993), as well as the growth of other
benthic algae that take advantage of the improved
light climate (Sand-Jensen & Søndergaard, 1981).
This may largely compensate for a reduced phyto-
plankton production (Liboriussen & Jeppesen, 2003;
Vadeboncoeur et al., 2003) and thereby maintain low
levels of CO
2
in the water. For example, Liboriussen
et al. (2011) and Jeppesen et al. (unpublished data)
found only minor variation in ecosystem production
and respiration for a number of 1-m deep ponds with
dominance either by macrophytes or phytoplankton
and benthic algae. Similar evidence was derived from
high-frequency oxygen measurements in L. Væng
conducted before, during and after the second
biomanipulation (Jeppesen et al., 2012). Net produc-
tion (March 1–Nov 15) ranged between 0.49 and
0.52 mg O
2
l
-1
day
-1
in 2007–2008 before restora-
tion and became negative in 2009 (-0.65 mg O
2
l
-1
day
-1
) when the lake was in a clear state without
macrophytes, but increased to 0.23 mg O
2
l
-1
day
-1
in 2010, coinciding with extensive growth of the
submerged macrophyte E. canadensis. In contrast,
comparison of two German lakes showed higher
gross system production in a macrophyte- compared
with a phytoplankton-dominated state (Brothers
et al., 2013).
In deep lakes, however, macrophyte and benthic
algal production may be limited to near-shore areas,
and total primary production typically decreases when
the phytoplankton production in the water declines, for
instance after restoration (Vadeboncoeur et al., 2008;
Genkai-Kato et al., 2012). Accordingly, the effect of
restoration by biomanipulation will lead to different
CO
2
emissions depending on lake depth, promoting
larger CO
2
efflux in deep lakes than in shallow systems
as the pelagic primary production will not be compen-
sated for by macrophytes or benthic production.
Table 3 Multiple regression (log-transformed data) with
AR(1) auto-regression included, relating CO
2
efflux (actual
value ?300 to account for negative values) to chlorophyll
aand macrophyte coverage (in % of lake area—PAC—one
added to account for zeroes) for annual data, summer (May 1–
Oct 1) and outside the growing season (Oct 1–Apr 1). Numbers
in parentheses are SD, N is number of years included, Pis the
significance level and R
2
the explained variance
Lake Intercept Chlorophyll aMacrophyte coverage (%, PAC) NP R
2
Annual
Faarup 8.19 (0.36) -0.51 (0.12) – 14 0.0008 0.65
Væng 7.92 (0.45) -0.24 (0.12) – 20 0.027 0.19
Engelsholm 7.93 (0.62) -0.33 (0.17) – 21 0.00006 0.43
Arreskov 6.75 (0.15) – -0.16 (0.05) 14 0.0018 0.47
Stigsholm 6.48 (0.19) – -0.14 (0.07) 20 0.021 0.20
May–September
Faarup 7.65 (0.45) -0.44 (0.27) – 14 0.0013 0.50
Væng 5.54 (0.37) -0.23 (0.11) – 19 0.006 0.06
Engelsholm 8.03 (0.48) -0.50 (0.10) 21 0.00004 0.45
Arreskov 7.59 (0.48) -0.29 (0.13) -0.21 (0.07) 14 0.0006 0.57
Stigsholm 6.21 (0.22) -0.29 (0.10) -0.39 (0.07) 20 0.00004 0.41
October–April
Faarup 8.30 (0.47) -0.52 (0.19) – 14 0.0002 0.67
Væng 7.07 (0.12) – -0.11 (0.04) 19 0.0011 0.74
Engelsholm 7.05 (0.14) – – 21 0.0062 0.08
Arreskov 6.93 (0.13) – -0.14 (0.04) 14 0.0056 0.32
Stigsholm 6.39 (0.08) – – 23 0.96 0.88
40 Hydrobiologia (2016) 778:33–44
123

We conclude that all the lakes studied were net-
releasers of CO
2
on an annual basis despite variations
in trophic webs. Eutrophication can decrease the
relative importance of external organic matter and
promote a higher autotrophic fixation of CO
2
(carbon
sink). It may, however, also promote a higher respi-
ration (carbon source) and increased release of other
greenhouse gases with a greater warming potential
such as methane (CH
4
) (Bastviken et al., 2008) and
N
2
O (Huttunen et al., 2003) from anoxic waters and
sediments. Lake restoration by nutrient loading reduc-
tion (Trolle et al., 2012) and/or biomanipulation by
fish removal (this study) may, however, result in a
further increase in the efflux of CO
2
, at least in the
short term. So, solving one problem (eutrophication)
partly enhances, at least temporarily, another (emis-
sion of carbon, further promoting climate warming,
depending on the balance among emissions of differ-
ent greenhouse gases).
When macrophytes become abundant, as they may
in shallow lakes of intermediate and high trophic state,
a shift back to enhanced retention of CO
2
is to be
expected. Submerged macrophytes have a strong
positive impact on biodiversity (Jeppesen et al.,
2000; Declerck et al., 2005; Muylaert et al., 2010)
and at least in temperate lakes also on water clarity
(Moss, 1990; Scheffer et al., 1993). The data presented
here suggest that plants also play a key role in reducing
the CO
2
efflux from lakes, as is also evidenced in a
mesocosm study by Davidson et al. (2015).
Restoration of shallow lakes is not always accom-
panied by the establishment of submerged macro-
phytes despite sufficient water clarity (Lauridsen et al.,
2003, Jeppesen et al., 2012). Delays in (re)colonisation
have been attributed to lack of sufficient propagules
and low dispersal potential or limited connection with
other aquatic systems acting as sources (Strand &
Weisner, 2001), herbivory at an early stage of
colonisation by waterfowl (Lauridsen et al., 1993;
Søndergaard et al., 1996; Marklund et al., 2002), fish
(Prejs, 1984) or crayfish (Rodrı
´guez-Gallego et al.,
2004) and adverse impacts of herbicides from agri-
cultural catchments. Active transplantation of macro-
phytes has been recommended when macrophytes are
Fig. 2 Conceptual overview of the expected behaviour of
shallow eutrophic lakes as sources or sinks of CO
2
, from a turbid
phytoplankton-dominated state to a clear water submerged
macrophyte-dominated state through biomanipulation and
colonisation of invasive mussels (such as zebra mussels). The
width of the arrows indicates the relative strength of CO
2
efflux
and influx
Hydrobiologia (2016) 778:33–44 41
123

not easily established naturally in order to stabilise the
clear water state and enhance ecosystem quality and
biodiversity (Moss, 1990; Scheffer et al., 1993; Moss
et al., 1996; Jeppesen et al., 1998).
Based on our study, a further argument in favour of
such restoration measures is that macrophyte estab-
lishment will also reduce greenhouse gas emissions, at
least of CO
2
and perhaps also of CH
4
(Davidson et al.,
2015).
The overwhelming majority of the world’s lakes are
small and shallow (Downing et al., 2006; Verpoorter
et al., 2014) and many of them will require restoration
in the decades to come given the increasing human
demands for fresh water (Dudgeon et al., 2006;
Vo
¨ro
¨smarty et al., 2010), but restoration measures
may have profound effects on the role of lakes within
the global carbon cycle.
Acknowledgments We are grateful to CRES (Centre for
Regional Change in the Earth System), the MARS project
(Managing Aquatic ecosystems and water Resources under
multiple Stress) funded under the 7th EU Framework
Programme, Theme 6 (Environment including Climate Change),
Contract No.: 603378 (http://www.mars-project.eu), CLEAR (a
Villum Kann Rasmussen Centre of Excellence project on lake
restoration) and CIRCE (Centre of Ecoinformatics Research in
Complexityin Ecology funded by the AU IDEAS programme) for
providing financial support. MM is supported by PEDECIBA,
SNI-ANII and the L
´Ore
´al-UNESCO for Women in Science
National Award.
References
Anderson, N. J., H. Bennion & A. F. Lotter, 2014. Lake
eutrophication and its implications for organic carbon
sequestration in Europe. Global Changes in Biology 20:
2741–2751.
Bade, D. L. & J. J. Cole, 2006. Impact of chemically enhanced
diffusion on dissolved inorganic carbon stable isotopes in a
fertilized lake. Journal of Geophysical Research 111.
doi:10.1029/2004JC002684.
Barros, N., J. J. Cole, L. J. Tranvik, Y. T. Prairie, D. Bastviken,
V. L. M. Huszar, P. del Giorgio & F. Roland, 2011. Carbon
emission from hydroelectric reservoirs linked to reservoir
age and latitude. Nature Geoscience 4: 593–596.
Bastviken, D., J. J. Cole, M. L. Pace & L. Tranvik, 2008. Fates of
methane from different lake habitats: connecting whole-
lake budgets and CH
4
emissions. Journal of Geophysical
Research 113: G02024.
Bjerring, R., L.S. Johansson, M. Søndergaard, E. Jeppesen, T.L.
Lauridsen, A. Kjeldgaard, L. Sortkjær, J. Windolf & J.
Bøgestrand, 2013. Søer 2013. NOVANA. Aarhus Univer-
sitet, DCE – Nationalt Center for Miljø og Energi, 84 s. -
Videnskabelig rapport fra DCE - Nationalt Center for Miljø
og Energi nr. 76. (in Danish)
Boll, T., L. S. Johansson, T. L. Lauridsen, F. Landkildehus, M.
Søndergaard, F. Ø. Andersen & E. Jeppesen, 2012. Chan-
ges in benthic macroinvertebrate community and lake
isotope (C, N) signals following a shift from clear to turbid
water in a shallow lake. Hydrobiologia 686: 135–145.
Brothers, S. M., S. Hilt, S. Meyer & J. Ko
¨hler, 2013. Plant
community structure determines primary productivity in
shallow, eutrophic lakes. Freshwater Biology 58:
2264–2276.
Carpenter, S. R., 1981. Submersed vegetation: an internal factor
in lake ecosystem succession. The American Naturalist
118: 372–383.
Carpenter, S. R., N. F. Caraco, D. L. Correll, R. W. Howarth, A.
N. Sharpley & V. H. Smith, 1998. Nonpoint pollution of
surface waters with phosphorus and nitrogen. Ecological
Applications 8: 559–568.
Cole, J. J., 2013. Freshwater ecosystems and the carbon cycle. In
Kinne, O. (ed.), Excellence in Ecology. Book 18. Interna-
tional Ecology Institute, Oldendorf/Luhe.
Cole, J. J., N. F. Caraco, G. W. Kling & T. K. Kratz, 1994.
Carbon dioxide supersaturation in the surface waters of
lakes. Science 265: 1568–1570.
Cole, J. J., M. L. Pace, S. R. Carpenter & J. F. Kitchell, 2000.
Persistence of net heterotrophy in lakes during nutrient
addition and food web manipulations. Limnology and
Oceanography 45: 1718–1730.
Cole, J. J., Y. T. Prairie, N. F. Caraco, W. H. McDowell, L.
J. Tranvik, R. G. Striegl, C. M. Duarte, P. Kortelainen, J.
A. Downing, J. J. Middelburg & J. Melack, 2007. Plumbing
the global carbon cycle: integrating inland waters into the
terrestrial carbon budget. Ecosystems 10: 171–184.
Davidson, T. A., J. Audet, J.-C. Svenning, T. L. Lauridsen, M.
Søndergaard, F. Landkildehus, S. E. Larsen & E. Jeppesen,
2015. Eutrophication effects on greenhouse gas fluxes from
shallow lake mesocosms override those of climate warm-
ing. Global Change Biology. doi:10.1111/gcb.13062.
Dean, W. E. & E. Gorham, 1998. Magnitude and significance of
carbon burial in lakes, reservoirs and peatlands. Geology
26: 535–538.
Declerck, S., J. Vandekerkhove, L. S. Johansson, K. Muylaert, J.
M. Conde-Porcuna, K. van der Gucht, T. Lauridsen, K.
Schwenk, G. Zwart, W. Rommens & J. Lopez-Ramos,
2005. Multi-group biodiversity in shallow lakes along
gradients of phosphorus and water plant cover. Ecology 86:
1905–1915.
de Kluijver, A., J. Ning, Z. Liu, E. Jeppesen & J. J. Middelburg,
2015. Macrophyte and periphyton carbon subsidies to
bacterioplankton and zooplankton in a shallow, eutrophic
lake in tropical China. Limnology and Oceanography 60:
375–385.
Downing, J. A., Y. T. Prairie, J. J. Cole, M. Duarte, L. J. Tranvik,
R. G. Striegl, W. H. McDowell, P. Kortelainen, N.
F. Caraco, J. M. Melack & J. J. Middelburg, 2006. The
global abundance and size distribution of lakes, ponds, and
impoundments. Limnology and Oceanography 51:
2388–2397.
Dudgeon, D., A. H. Arthington, M. O. Gessner, Z.-I. Kawabata,
D. J. Knowler, C. Le
´ve
ˆque, R. J. Naiman, A.-H. Prieur-
Richard, D. Soto, M. L. J. Stiassny & C. A. Sullivan, 2006.
42 Hydrobiologia (2016) 778:33–44
123

Freshwater biodiversity: importance, threats, status and
conservation challenges. Biological Reviews 81: 163–182.
Gao, J., Z. Liu & E. Jeppesen, 2014. Fish trophic structures but
not biomass changed after lake restoration by biomanipu-
lation in a tropical eutrophic lake. Hydrobiologia 724:
127–140.
Genkai-Kato, M., Y. Vadeboncoeur, L. Liboriussen & E.
Jeppesen, 2012. Benthic-pelagic coupling, regime shifts
and whole-lake primary production in shallow lakes.
Ecology 93: 619–631.
Higgins, S. N. & M. J. Vander Zanden, 2010. What a difference
a species makes: a meta-analysis of dreissenid mussel
impacts on freshwater ecosystems. Ecological Monographs
80: 179–196.
Huttunen, J. T., S. Juutinen, J. Alm, T. Larmola, T. Hammar, J.
Silvola & P. J. Martikainen, 2003. Fluxes of methane,
carbon dioxide and nitrous oxide in boreal lakes and
potential anthropogenic effects on the aquatic greenhouse
gas emissions. Chemosphere 52: 609–621.
Idrisi, N., E. L. Mills, L. G. Rudstam & D. J. Stewart, 2001.
Impact of zebra mussels (Dreissena polymorpha) on the
pelagic lower trophic levels of Oneida Lake, New York.
Candian Journal of Fisheries and Aquatic Sciences 58:
1430–1441.
Jeppesen, E., J. P. Jensen, M. Søndergaard, T. Lauridsen, F.
P. Møller & K. Sandby, 1998. Changes in nitrogen reten-
tion in shallow eutrophic lakes following a decline in
density of cyprinids. Archiv fu
¨r Hydrobiologie 142:
129–151.
Jeppesen, E., J. P. Jensen, M. Søndergaard, T. Lauridsen & F.
Landkildehus, 2000. Trophic structure, species richness
and biodiversity in Danish lakes: changes along a phos-
phorus gradient. Freshwater Biology 45: 201–213.
Jeppesen, E., M. Søndergaard, T. L. Lauridsen, T. A. Davidson,
Z. Liu & N. Mazzeo, 2012. Biomanipulation as a restoration
tool to combat eutrophication: recent advances and future
challenges. Advances in Ecological Research 47: 411–487.
Jones, J. I., 2005. The metabolic cost of bicarbonate use in the
submerged plant Elodea nuttallii. Aquatic Botany 83:
71–81.
Jones, J. I., K. Hardwick & J. W. Eaton, 1996. Diurnal carbon
restrictions on the photosynthesis of dense stands of Elodea
nuttallii (Planch.) St. John. Hydrobiologia 340: 11–16.
Jones, J. I., J. W. Eaton & K. Hardwick, 2000. The effect of
changing environmental variables in the surrounding water
on the physiology of Elodea nuttallii. Aquatic Botany 66:
115–129.
Karatayev, A. Y., D. K. Padilla, D. Minchin, D. Boltovskoy & L.
E. Burlakova, 2007. Changes in global economies and
trade: the potential spread of exotic freshwater bivalves.
Biological Invasions 9: 161–180.
Kidmose, J., B. Nilsson, P. Engesgaard, M. Frandsen, S. Karan,
F. Landkildehus, M. Søndergaard & E. Jeppesen, 2013.
Focused groundwater discharge of phosphorus to a
eutrophic seepage lake (Lake Væng, Denmark): implica-
tions for lake ecological state and restoration. Hydrogeol-
ogy Journal 2: 1787–1802.
Kling, G. W., G. W. Kipphut & M. C. Miller, 1992. The flux of
CO
2
and CH
4
from lakes and rivers in arctic Alaska.
Hydrobiologia 240: 23–36.
Kosten, S., F. Roland, D. M. L. Da Motta Marques, E. H. Van
Nes, N. Mazzeo, L. D. S. L. Sternberg, M. Scheffer & J.
J. Cole, 2010. Climate-dependent CO
2
emissions from
lakes. Global Biogeochemical Cycles 24: GB2007.
Kronvang, B., G. Ærtebjerg, R. Grant, P. Kristensen, M. Hov-
mand & J. Kirkegaard, 1993. Nation-wide Danish moni-
toring programme – state of the aquatic environment.
Ambio 22: 176–187.
Lauridsen, T., E. Jeppesen & F. Ø. Andersen, 1993. Coloniza-
tion of sub-merged macrophytes in shallow fish manipu-
lated lake Væng: Impact of sediment compo-si-tion and
birds grazing. Aquatic Botany 46: 1–15.
Lauridsen, T. L., E. Jeppesen & M. Søndergaard, 1994. Colo-
nization and succession of submerged macrophytes in
shallow Lake Væng during the first five years following
fish manipulation. Hydrobiologia 275(276): 233–242.
Lauridsen, T. L., J. P. Jensen, E. Jeppesen & M. Søndergaard,
2003. Response of submerged macrophytes in Danish lakes
to nutrient loading reductions and biomanipulation.
Hydrobiologia 506–509: 641–649.
Liboriussen, L. & E. Jeppesen, 2003. Temporal dynamics in
epipelic, pelagic and epiphytic algal production in a clear
and a turbid shallow lake. Freshwater Biology 48:
418–431.
Liboriussen, L., M. Søndergaard & E. Jeppesen (red.), 2007.
Sørestaurering i Danmark. Faglig rapport fra DMU 636:
del I 86 s ?del II 312 s. (in Danish)
Liboriussen, L., T. L. Lauridsen, M. Søndergaard, F. Land-
kildehus, M. Søndergaard & E. Jeppesen, 2011. Climate
warming effect on the seasonal dynamics in sediment
respiration in shallow lakes: an outdoor mesocosms
experiment. Freshwater Biology 56: 437–447.
Marklund, O., H. Sandsten, L.-A. Hansson & I. Blindow, 2002.
Effects of waterfowl and fish on submerged vegetation and
macroinvertebrates. Freshwater Biology 47: 2049–2059.
Mayer, C. M., R. A. Keats, L. G. Rudstam & E. L. Mill, 2002.
Scale-dependent effects of zebra mussels on benthic
invertebrates in a large eutrophic lake. Journal of the North
American Benthological Society 21: 616–633.
Moss, B., 1990. Engineering and biological approaches to the
restoration from eutrophication of shallow lakes in which
aquatic plant communities are important components.
Hydrobiologia 200(201): 367–377.
Moss, B., J. Madgwick & G. L. Phillips, 1996. A Guide to the
Restoration of Nutrient-Enriched Shallow Lakes. Broads
Authority & Environment Agency, Norwich.
Moss, B., S. Kosten, M. Meerhoff, R. W. Battarbee, E. Jeppesen,
N. Mazzeo, K. Havens, G. Lacerot, Z. Liu, L. De Meester,
H. Paerl & M. Scheffer, 2011. Allied attack: climate
change and nutrient pollution. Inland Waters 1: 101–105.
Muylaert, K., C. Pe
´rez-Martı
´nez, P. Sa
´nchez-Castillo, T.
L. Lauridsen, M. Vanderstukken, S. A. J. Declerck, K. Van
der Gucht, J. M. Conde-Porcuna, E. Jeppesen, L. De
Meester & W. Vyverman, 2010. Influence of nutrients,
submerged macrophytes and zooplankton grazing on
phytoplankton biomass and diversity along a latitudinal
gradient in Europe. Hydrobiologia 653: 79–90.
Nielsen, A., D. Trolle, R. Bjerring, M. Søndergaard, J. E. Olesen
& E. Jeppesen, 2014. Effects of climate and nutrient load
on the water quality of shallow lakes assessed through
Hydrobiologia (2016) 778:33–44 43
123

ensemble runs by PCLake. Ecological Applications 24:
1926–1944.
Peterson, B. J. & B. Fry, 1987. Stable isotopes in ecosystem
studies. Annual Review of Ecology and Systematics 18:
293–320.
Prejs, A., 1984. Herbivory by temperate freshwater fishes and its
consequences. Environmental Biology of Fishes 10:
281–296.
Provoost, P., S. van Heuven, K. Soetaert, R. Laane & J.
J. Middelburg, 2010. Long-term record of pH in the Dutch
coastal zone: a major role for eutrophication-induced
changes. Biogeosciences Discussions 7: 4127–4152.
Raymond, P. A., J. Hartmann, R. Lauerwald, S. Sobek, C.
McDonald, M. Hoover, D. Butman, R. Striegl, E. Mayorga,
C. Humborg, P. Kortelainen, H. Durr, M. Meybeck, P.
Ciais & P. Guth, 2013. Global carbon dioxide emissions
from inland waters. Nature 503: 355–359.
Rodrı
´guez-Gallego, L., N. Mazzeo, J. Gorga, M. Meerhoff, J.
Clemente, C. Kruk, F. Scasso, G. Lacerot, J. Garcı
´a&F.
Quintans, 2004. Effects of an artificial wetland with free-
floating plants on the restoration of a hypertrophic sub-
tropical lake. Lake and Reservoir Management 9: 203–215.
Sandby, K. S. & J. Hansen, 2007. Lake Arreskov. In: Libo-
riussen, L., M. Søndergaard, E. Jeppesen (eds) Sørestau-
rering i Danmark. Report from NERI no. 636 (in Danish).
Scheffer, M., S. H. Hosper, M. L. Meijer, B. Moss & E.
Jeppesen, 1993. Alternative equilibria in shallow lakes.
Trends in Ecology and Evolution 8: 275–279.
Stallard, R. F., 1998. Terrestrial sedimentation and the carbon
cycle: coupling weathering and erosion to carbon burial.
Global Biogeochemical Cycles 12: 231–257.
Stewart, T. W., J. G. Miner & R. L. Lowe, 1998. Quantifying
mechanisms for zebra mussel effects on benthic macroin-
vertebrates: organic matter production and shell-generated
habitat. Journal of the North American Benthological
Society 17: 81–94.
Stumm, W. & J. J. Morgan, 1996. Aquatic Chemistry. John
Wiley, New York.
Strand, J. A. & S. E. B. Weisner, 2001. Dynamics of submerged
macrophyte populations in response to biomanipulation.
Freshwater Biology 46: 1397–1408.
Strayer, D. L., N. F. Caraco, J. J. Cole, S. Findlay & M. L. Pace,
1999. Transformation of freshwater ecosystems by
bivalves – a case study of zebra mussels in the Hudson
River. Bioscience 49: 19–27.
Søndergaard, M., E. Jeppesen, E. Mortensen, E. Dall, P. Kris-
tensen & O. Sortkjær, 1990. Phytoplankton biomass
reduction after planktivorous fish reduction in a shallow,
eutrophic lake: a combined effect of reduced internal
P-loading and increased zooplankton grazing. Hydrobi-
ologia 200(201): 229–240.
Søndergaard, M., L. Olufsen, T. Lauridsen, E. Jeppesen & T.
Vindbæk Madsen, 1996. The impact of grazing waterfowl
on submerged macrophytes: in situ experiments in a shal-
low eutrophic lake. Aquatic Botany 53: 73–84.
Søndergaard, M., T. L. Lauridsen, E. Jeppesen & L. Bruun,
1998. Macrophyte-Waterfowl Interactions: Tracking a
Variable Resource and the Impact of Herbivory on Plant
Growth. In Jeppesen, E., M. Søndergaard, M. Søndergaard
& K. Christoffersen (eds.), The Structuring Role of Sub-
merged Macrophytes in Lakes, Vol. 131., Ecological
Studies Series Springer, New York: 298–306.
Talling, J. F., 1976. The depletion of carbon dioxide from lake
water by phytoplankton. Journal of Ecology 64: 79–121.
Tranvik, L. J., J. A. Downing, J. B. Cotner, S. A. Loiselle, R.
G. Striegl, T. J. Ballatore, P. Dillon, L. B. Knoll, T. Kutser,
S. Larsen, I. Laurion, D. M. Leech, S. L. McAllister, D.
M. McKnight, J. Melack, E. Overholt, J. A. Porter, Y.
T. Prairie, W. H. Renwick, F. Roland, B. S. Sherman, D.
W. Schindler, S. Sobek, A. Tremblay, M. J. Vanni, A.
M. Verschoor, E. von Wachenfeldt & G. Weyhenmeyer,
2009. Lakes and reservoirs as regulators of carbon cycling
and climate. Limnology and Oceanography 54: 2298–2314.
Trolle, D., P. A. Staehr, T. A. Davidson, R. Bjerring, T.
L. Lauridsen, M. Søndergaard & E. Jeppesen, 2012. Sea-
sonal dynamics of CO
2
flux across the surface of shallow
temperate lakes of contrasting trophic status, and the
response to nutrient load reduction. Ecosystems 15:
336–347.
Vadeboncoeur, Y., E. Jeppesen, M. J. Vander Zanden, H.
H. Schierup, K. Christoffersen & D. Lodge, 2003. From
Greenland to green lakes: cultural eutrophication and the
loss of benthic pathways in lakes. Limnology and
Oceanography 48: 1408–1418.
Vadeboncoeur, Y., G. Peterson, M. J. Vander Zanden & J. Kalff,
2008. Benthic algal productivity across lake size gradients:
interactions among morphometry, nutrients, and light.
Ecology 89: 2542–2552.
Verpoorter, C., T. Kutser, D. A. Seekell & L. J. Tranvik, 2014. A
global inventory of lakes based on high-resolution satellite
imagery. Geophysical Research Letters 41: 6396–6402.
Vo
¨ro
¨smarty, C. J., P. B. McIntyre, M. O. Gessner, D. Dudgeon,
A. Prusevich, P. Green, S. Glidden, S. E. Bunn, C. A. Sul-
livan, C. Reidy Liermann & P. M. Davies, 2010. Global
threats to human water security and river biodiversity.
Nature 467: 555–561.
Wetzel, R. G., 2001. Limnology: Lake and River Ecosystems.
Academic Press, San Diego.
44 Hydrobiologia (2016) 778:33–44
123