Response of Submerged Macrophyte Communities to External and Internal Restoration Measures in North Temperate Shallow Lakes

Abstract

Submerged macrophytes play a key role in north temperate shallow lakes by stabilizing clear-water conditions. Eutrophication has resulted in macrophyte loss and shifts to turbid conditions in many lakes. Considerable efforts have been devoted to shallow lake restoration in many countries, but long-term success depends on a stable recovery of submerged macrophytes. However, recovery patterns vary widely and remain to be fully understood. We hypothesize that reduced external nutrient loading leads to an intermediate recovery state with clear spring and turbid summer conditions similar to the pattern described for eutrophication. In contrast, lake internal restoration measures can result in transient clear-water conditions both in spring and summer and reversals to turbid conditions. Furthermore, we hypothesize that these contrasting restoration measures result in different macrophyte species composition, with added implications for seasonal dynamics due to differences in plant traits. To test these hypotheses, we analyzed data on water quality and submerged macrophytes from 49 north temperate shallow lakes that were in a turbid state and subjected to restoration measures. To study the dynamics of macrophytes during nutrient load reduction, we adapted the ecosystem model PCLake. Our survey and model simulations revealed the existence of an intermediate recovery state upon reduced external nutrient loading, characterized by spring clear-water phases and turbid summers, whereas internal lake restoration measures often resulted in clear-water conditions in spring and summer with returns to turbid conditions after some years. External and internal lake restoration measures resulted in different macrophyte communities. The intermediate recovery state following reduced nutrient loading is characterized by a few macrophyte species (mainly pondweeds) that can resist wave action allowing survival in shallow areas, germinate early in spring, have energy-rich vegetative propagules facilitating rapid initial growth and that can complete their life cycle by early summer. Later in the growing season these plants are, according to our simulations, outcompeted by periphyton, leading to late-summer phytoplankton blooms. Internal lake restoration measures often coincide with a rapid but transient colonization by hornworts, waterweeds or charophytes. Stable clear-water conditions and a diverse macrophyte flora only occurred decades after external nutrient load reduction or when measures were combined.

Full text

ORIGINAL RESEARCH
published: 19 February 2018
doi: 10.3389/fpls.2018.00194
Frontiers in Plant Science | www.frontiersin.org 1February 2018 | Volume 9 | Article 194
Edited by:
Janne Alahuhta,
University of Oulu, Finland
Reviewed by:
Rebecca Lester,
Deakin University, Australia
Ludwig Triest,
Vrije Universiteit Brussel, Belgium
*Correspondence:
Sabine Hilt
hilt@igb-berlin.de
Specialty section:
This article was submitted to
Functional Plant Ecology,
a section of the journal
Frontiers in Plant Science
Received: 28 September 2017
Accepted: 01 February 2018
Published: 19 February 2018
Citation:
Hilt S, Alirangues Nuñez MM,
Bakker ES, Blindow I, Davidson TA,
Gillefalk M, Hansson L-A, Janse JH,
Janssen ABG, Jeppesen E, Kabus T,
Kelly A, Köhler J, Lauridsen TL,
Mooij WM, Noordhuis R, Phillips G,
Rücker J, Schuster H-H,
Søndergaard M, Teurlincx S,
van de Weyer K, van Donk E,
Waterstraat A, Willby N and Sayer CD
(2018) Response of Submerged
Macrophyte Communities to External
and Internal Restoration Measures in
North Temperate Shallow Lakes.
Front. Plant Sci. 9:194.
doi: 10.3389/fpls.2018.00194
Response of Submerged Macrophyte
Communities to External and Internal
Restoration Measures in North
Temperate Shallow Lakes
Sabine Hilt 1
*, Marta M. Alirangues Nuñez 1, Elisabeth S. Bakker 2, Irmgard Blindow 3,
Thomas A. Davidson 4, Mikael Gillefalk 1, Lars-Anders Hansson 5, Jan H. Janse 2,6 ,
Annette B. G. Janssen 2,7 , Erik Jeppesen 4,8 , Timm Kabus 9, Andrea Kelly 10, Jan Köhler 1,
Torben L. Lauridsen 4,8, Wolf M. Mooij 2, 11, Ruurd Noordhuis 12 , Geoff Phillips 13 ,
Jacqueline Rücker 14, Hans-Heinrich Schuster 15 , Martin Søndergaard 4,8 , Sven Teurlincx2,
Klaus van de Weyer 16, Ellen van Donk 2, Arno Waterstraat 17, Nigel Willby 13 and
Carl D. Sayer 18
1Department of Ecosystem Research, Leibniz-Institute of Freshwater Ecology and Inland Fisheries, Berlin, Germany,
2Departmnet of Aquatic Ecology, Netherlands Institute of Ecology (NIOO-KNAW), Wageningen, Netherlands, 3Biological
Station of Hiddensee, University of Greifswald, Greifswald, Germany, 4Department of Bioscience, Aarhus University,
Silkeborg, Denmark, 5Department of Biology, Lund University, Lund, Sweden, 6Netherlands Environmental Assessment
Agency (PBL), Den Haag, Netherlands, 7Water Systems and Global Change Group, Wageningen University and Research,
Wageningen, Netherlands, 8Sino-Danish Centre for Education and Research, University of Chinese Academy of Sciences,
Beijing, China, 9Institute of Applied Freshwater Ecology, Seddiner See, Germany, 10 Broads Authority, Norwich,
United Kingdom, 11 Department of Aquatic Ecology and Water Quality Management, Wageningen University and Research,
Wageningen, Netherlands, 12 Deltares, Delft, Netherlands, 13 Biological and Environmental Sciences, University of Stirling,
Stirling, United Kingdom, 14 Department of Freshwater Conservation, Brandenburg University of Technology
Cottbus-Senftenberg, Senftenberg, Germany, 15 Niedersächsischer Landesbetrieb für Wasserwirtschaft, Küsten- und
Naturschutz, Sulingen, Germany, 16 Lanaplan, Nettetal, Germany, 17 Gesellschaft für Naturschutz und Landschaftsökologie,
Kratzeburg, Germany, 18 Department of Geography, Environmental Change Research Centre, University College London,
London, United Kingdom
Submerged macrophytes play a key role in north temperate shallow lakes by stabilizing
clear-water conditions. Eutrophication has resulted in macrophyte loss and shifts to
turbid conditions in many lakes. Considerable efforts have been devoted to shallow lake
restoration in many countries, but long-term success depends on a stable recovery
of submerged macrophytes. However, recovery patterns vary widely and remain to
be fully understood. We hypothesize that reduced external nutrient loading leads
to an intermediate recovery state with clear spring and turbid summer conditions
similar to the pattern described for eutrophication. In contrast, lake internal restoration
measures can result in transient clear-water conditions both in spring and summer
and reversals to turbid conditions. Furthermore, we hypothesize that these contrasting
restoration measures result in different macrophyte species composition, with added
implications for seasonal dynamics due to differences in plant traits. To test these
hypotheses, we analyzed data on water quality and submerged macrophytes from 49
north temperate shallow lakes that were in a turbid state and subjected to restoration
measures. To study the dynamics of macrophytes during nutrient load reduction, we
adapted the ecosystem model PCLake. Our survey and model simulations revealed
the existence of an intermediate recovery state upon reduced external nutrient loading,
characterized by spring clear-water phases and turbid summers, whereas internal lake

Hilt et al. Response of Macrophytes to Restoration
restoration measures often resulted in clear-water conditions in spring and summer
with returns to turbid conditions after some years. External and internal lake restoration
measures resulted in different macrophyte communities. The intermediate recovery state
following reduced nutrient loading is characterized by a few macrophyte species (mainly
pondweeds) that can resist wave action allowing survival in shallow areas, germinate early
in spring, have energy-rich vegetative propagules facilitating rapid initial growth and that
can complete their life cycle by early summer. Later in the growing season these plants
are, according to our simulations, outcompeted by periphyton, leading to late-summer
phytoplankton blooms. Internal lake restoration measures often coincide with a rapid
but transient colonization by hornworts, waterweeds or charophytes. Stable clear-water
conditions and a diverse macrophyte flora only occurred decades after external nutrient
load reduction or when measures were combined.
Keywords: aquatic plants, biomanipulation, eutrophication, lake restoration. nutrient load reduction, PCLake, plant
traits, regime shift
INTRODUCTION
Shallow lakes are the most abundant freshwater ecosystems on
earth (Verpoorter et al., 2014). In their pristine state, they are
often characterized by abundant submerged vegetation which can
stabilize clear-water conditions (Scheffer et al., 1993) and plays a
key role in the functioning of the ecosystem (Hilt et al., 2017).
Several mechanisms contribute to a positive feedback between
macrophytes and clear water conditions. As a consequence,
shallow lakes are resistant to increasing nutrient loading up
to a critical threshold, above which their macrophytes collapse
and the lakes shift into a turbid, phytoplankton-dominated state
(Scheffer et al., 1993). In recent centuries, excessive nutrient
loading has resulted in a loss of macrophytes and shift to this
turbid state in many temperate shallow lakes (e.g., Körner, 2002;
Phillips et al., 2016).
Sayer et al. (2010a) suggested a typical pattern of lake
macrophyte loss, defining a so-called “crashing” state lying
between the stable clear-water state featuring a diverse plant
community and the final turbid state lacking in macrophytes.
This crashing state is characterized by the occurrence of only a
few macrophyte species that can complete their life cycle during
clear-water conditions in spring and early summer while later
in summer, cyanobacteria blooms often occur. Eventually, the
remaining macrophyte stands are also lost and give way to year-
round phytoplankton dominance (Sayer et al., 2010a,b, 2016).
Under these conditions, several ecosystem functions and services
deteriorate, including biodiversity support, nutrient retention,
provision of water of drinking or swimming quality (Hilt et al.,
2017).
Hence, considerable efforts and financial resources have been
devoted to the restoration of shallow lakes in many countries
in recent decades (Jeppesen et al., 2005). The success of lake
restoration in the long-term depends critically on the stable
recovery of submerged macrophytes (Hilt et al., 2006). However,
the turbid state is stabilized by feedback mechanisms that can
prevent macrophyte re-colonization even at reduced nutrient
loading. In theory, only the reduction of nutrient levels below a
critical threshold or a significant reduction in the abundance of
planktivorous and benthivorous fish (e.g., by biomanipulation or
natural fish kills) will lead to a recovery of clear-water conditions
and a return of macrophytes (Scheffer et al., 1993). In practice,
reductions in the external nutrient load to shallow lakes often fail
to deliver macrophyte recovery (Jeppesen et al., 2005). Similarly,
biomanipulation of the fish community in turbid shallow lakes
has produced variable effects on macrophytes in shallow lakes
(Hansson et al., 1998; Bergman et al., 1999; Søndergaard et al.,
2008; Jeppesen et al., 2012; Bernes et al., 2015; Sayer et al., 2016).
Overall, the response of macrophyte communities to different
types of lake restoration measures remains to be fully understood
(Jeppesen et al., 2005; Bakker et al., 2013).
We hypothesize that (1) external lake restoration measures
leading to nutrient load reduction in turbid temperate shallow
lakes result in macrophyte re-establishment in a reversed
sequence to the one described by Sayer et al. (2010a,b) for
advancing eutrophication. An intermediate recovery state should
occur where the water is clear in spring but dominated by
phytoplankton and thus turbid in late summer, until, eventually,
seasonally stable conditions characterized by high water clarity in
both spring and summer would dominate (Figure 1). In contrast,
lake internal measures such as biomanipulation or phosphorus
precipitation are expected to result in transient clear-water
conditions in spring and summer if either zooplankton is
sufficiently released from fish predation or internal phosphorus
loading from sediments is reduced enough to control summer
phytoplankton. Such conditions are supposed to occur only
temporarily in the absence of additional external nutrient load
reduction (Figure 1). We hypothesize that (2) these contrasting
types of restoration measures result in different macrophyte
community composition and seasonal patterns in plant
abundance. Specific macrophyte communities with short growth
seasons should dominate during the intermediate recovery
state following external nutrient loading reduction, while
species with longer growing season requirements are predicted
to temporarily establish following lake-internal measures
(Figure 1). The establishment of stable clear conditions with a
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Hilt et al. Response of Macrophytes to Restoration
FIGURE 1 | Response patterns of turbid north temperate shallow lakes to different restoration measures: (1) External restoration measures (reduction of external
nutrient loading) are expected to lead to an intermediate recovery state with clear-water conditions in spring and turbid water in summer and specific macrophyte
communities with short growth seasons and eventually stable clear conditions with a diverse macrophyte flora if nutrient loading is reduced sufficiently or additional
internal measures are applied (reversed order as suggested for eutrophication by Sayer et al., 2010a,b). Thresholds in phosphorus (P) loading are based on
simulations using PCLake (see Figure 5). (2) Lake-internal measures (biomanipulation, sediment suction dredging) leading to unstable clear-water conditions with
specific macrophyte communities that may collapse resulting in a shift back to turbid conditions unless nutrient loading is reduced, or (3) a combination of external
and internal restoration leading to stable clear-water conditions with an abundant and diverse macrophyte community.
diverse macrophyte community is thus assumed to require both,
external and internal measures (Figure 1).
To test these hypotheses, we analyse existing data on the water
quality and submerged macrophytes of 49 temperate shallow
lakes that had deteriorated to a turbid state and subsequently
were subject to either external or internal restoration measures
or both. In addition, we use an adapted version of the ecosystem
model PCLake (Janse et al., 2008) to simulate the response of
water clarity and macrophyte biomass to external nutrient load
reduction and to detect any thresholds in nutrient loading for
macrophyte recovery. Traits of the typical macrophyte species
found after external and internal lake restoration measures
are compared to provide a mechanistic understanding of the
observed re-colonization patterns.
MATERIALS AND METHODS
Literature and Data Search on Macrophyte
Species Recovery
We started our literature review with the 22 shallow lakes
described in detail in the study by Jeppesen et al. (2005) on the
response of lakes to reduced external nutrient loading. However,
only two of these 22 lakes were turbid before the nutrient
load reduction (Müggelsee, Veluwemeer) and both showed an
increase in macrophyte coverage following the intervention. The
rest showed no change, macrophyte declines or lacked suitable
data (Jeppesen et al., 2005). Therefore, we searched for more
examples in published and unpublished studies on lakes in
Germany, The Netherlands, Denmark, southern Sweden and
UK where shallow lakes are abundant and experience similar
eutrophication problems and climatic conditions. We selected
lakes that had lost most of their submerged macrophytes during
a turbid phase and subsequently had been subjected to either
external restoration via nutrient load reduction (summarized
in Table 1), or internal restoration using biomanipulation or
sediment dredging (Table 2). We did not carry out a full
systematic review of available data, but instead focussed on
known lakes within the research network of the authors where
at least partial recovery through restoration measures was
evident.
For all lakes, we retrieved information from the turbid
period and its macrophyte assemblage, macrophyte composition
after external/internal restoration, total phosphorus (TP)
concentrations in the water and Secchi depth in spring (April-
June) and summer (July–September) using published studies or
questionnaire responses provided by co-authors. As is commonly
the case, data on TP concentrations and Secchi depth were very
diverse, ranging from multi-year weekly measurements to
single values. The data were merged into a single value for
each season and lake by using means (Figures 2A,B) and
raw data are available as a supplement. We also analyzed
the occurrence of dominant macrophyte species during the
recovery period. To visualize potentially typical recovery
patterns, the long-term changes in nutrient concentrations,
water transparency and macrophyte occurrence are shown
in more detail for three lakes restored only by reductions in
external nutrient loading (Müggelsee, Veluwemeer, Eemmeer,
for details see no. 1, 19 and 20 in Table 1) and for three lakes
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Hilt et al. Response of Macrophytes to Restoration
TABLE 1 | The response of north temperate shallow lakes in Germany (DE), United Kingdom (UK), The Netherlands (NL), and Denmark (DK) to external nutrient load reduction (x: data on Secchi disk transparency and
total phosphorus concentrations were available and used in Figure 2A).
No. Lake Country Size
(ha)
Depth
(max/mean)
(m)
Measures of nutrient load
reduction
Period of turbid
conditions, major
remaining macrophyte
species and coverage (%
lake area)
Period of intermediate recovery
conditions, major macrophyte
species and coverage (% lake
area)
Period of stable
clear conditions
and macrophyte
species
Data
Figure 2A
References
1 Großer
Müggelsee
DE 750 8/4.9 Since 1989: improved
wastewater treatment in
catchment
1970–1989
Sparse stands of
P. pectinatus
1990–2013
P. pectinatus, P. perfoliatus,
P. crispus, Z. palustris
ca. 5% (Figure 4)
2014-now
P. pectinatus,
P. perfoliatus,
Najas marina,
E. nuttallii
ca. 25% (Figure 4)
xHilt et al., 2013, S. Hilt
unpubl. data
2 Großer
Wannsee
282 9.8/5.5 Nutrient load reduction in
catchment
?-ongoing
P. pectinatus, P. crispus
<5%
not yet reached Hilt and Grünert, 2008;
Van de Weyer, 2011
3 Galen-becker
See
590 1.8/0.8 Since 1972: reduction of P
loading by treatment in
upstream reservoir, since
2007 by regulation of whole
water supply to lake
1995–2002 2003–2006
P. pectinatus
2008–now
Chara contraria,
C. tomentosa,
C. globularis,
C. intermedia,
Nitellopsis obtusa,
N. marina f.
intermedia,
P. pectinatus,
Z. palustris
xWaterstraat, 2008,
unpubl. data
4 Dümmer 1350 4/1.1 Reduction of P load by
wastewater treatment plants
by 95%
1960–2011
P. pectinatus (rare)
2011–ongoing
P. pectinatus, P. crispus, Z. palustris,
P. pusillus
25-50%
not yet reached x Blüml et al., 2008;
Schuster, unpubl. data
5 Schwielow-
see
786 9.1/2.8 Nutrient load reduction in
catchment area (River Havel)
?−2005 2006–now
P. pectinatus
not yet reached x Kabus et al., 2007
6 Wuster-witzer
See
172 9.2/3.4 Nutrient load reduction in
catchment area
?−2005 2006–ongoing
P. pectinatus
not yet reached x Kabus et al., 2007
7 Grimnitz-see 777 10.3/4.5 Since 1994: sewage
treatment plant in operation
1970–1990
P. pectinatus
1991–?
P. pectinatus
?–now
2008: C. contraria,
N. obtusa,
P. pectinatus,
P. perfoliatus,
C. demersum, L.
trisulca
2001: C. contraria,
C. globularis,
N. obtusa,
P. pectinatus
xMauersberger and
Mauersberger, 1996;
Gervais et al., 1999;
Kabus and
Mauersberger, 2011
(Continued)
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Hilt et al. Response of Macrophytes to Restoration
TABLE 1 | Continued
No. Lake Country Size
(ha)
Depth
(max/mean)
(m)
Measures of nutrient load
reduction
Period of turbid
conditions, major
remaining macrophyte
species and coverage (%
lake area)
Period of intermediate recovery
conditions, major macrophyte
species and coverage (% lake
area)
Period of stable
clear conditions
and macrophyte
species
Data
Figure 2A
References
8 Wardersee 357 10.8/3.7 Wastewater treatment
plants
? 1996–2006
P. pectinatus, P. perfoliatus,
P. crispus, Z. palustris
? x Landesamt für Natur
und Umwelt des
Landes
Schleswig-Holstein,
1997; Heinzel and
Martin, 2006
9 Hemmels-
dorfer
See
450 6 (northern
part)
1998: P load reduction ?-1978-? ?-2006-?
P. pectinatus, Z.palustris,
P. perfoliatus, R. circinatus
?Heinzel and Martin,
2006
10 Großer
Varchen-tiner
See
182 1.7/? Lowered nutrient input from
agricultural catchment since
1990
?–? ?-2012-?
P. pectinatus,
M. spicatum
90%
2016
M. spicatum,
Najas marina ssp.
intermedia
x Kabus unpubl. data
11 Großer
Dambecker
See
94 2.1/0.8 Lowered nutrient input from
agricultural catchment since
1990
?- 2007
few P. pectinatus
2010-ongoing
P. pectinatus,
C. globularis
70%
Not yet reached x Kabus unpubl. data
12 Langer See 130 3.8/2.2 Since 1990: catchment
restoration, improved
wastewater treatment
?-1997-2001-? ?-2011-2015-ongoing
M. spicatum, C. demersum,
N. marina
not yet reached x Rücker et al., 2015,
unpubl. data
13 Nonnensee 76 2.2/? Re-flooded area (mid
1990’s)
?- 2012
very few
P. pusillus
2012, 2016
P. pectinatus,
M. spicatum, Ceratophyllum spp.
not yet reached Kabus unpubl. data
14 De Wittsee 24 2.1/1.4 Improvement of wastewater
treatment
1970-2009
Nuphar lutea
2009-ongoing
E. nuttallii, Lemna
not yet reached x Van de Weyer, unpubl.
data
15 Steinhuder
Meer
3000 2.9/1.35 Improvement of wastewater
treatment
1960-1998 none 1999-2001:
2002-03: E. nuttallii
2003-08: P. perfoliatus, P. crispus
2009-?: shift back to turbid
not yet reached Hussner et al., 2014
16 Felbrigg Lake UK 2.7 1.3/0.9 Creation of pre-lake wetland
in 2012 resulting in N-limited
conditions. Cormorant
predation on rudd
1960-2013
P. pectinatus, P. pusillus,
P. crispus, Z. palustris
2014-ongoing
C. demersum, Chara spp.
P. pectinatus, P. crispus, P. pusillus
unclear whether
already reached
Sayer et al., 2010a,b;
Sayer et al. unpubl.
data
17 Barton Broad 75 2/1 Progressive increase in
number of effluents with P
removal
1974-1990
none
1990-2000
C. demersum, P. crispus
<1%
reached after
sediment removal
1996 (Table 2, no.
41)
xPhillips et al., 2005,
2015
(Continued)
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Hilt et al. Response of Macrophytes to Restoration
TABLE 1 | Continued
No. Lake Country Size
(ha)
Depth
(max/mean)
(m)
Measures of nutrient load
reduction
Period of turbid
conditions, major
remaining macrophyte
species and coverage (%
lake area)
Period of intermediate recovery
conditions, major macrophyte
species and coverage (% lake
area)
Period of stable
clear conditions
and macrophyte
species
Data
Figure 2A
References
18 Wolderwijd NL 2650 2.5/1.5 1982-89: Flushing with
nutrient-poor water
1969-1975
P. pectinatus, P. perfoliatus
(Figure 5)
1976-1995
P. pectinatus,
P. perfoliatus, P. pusillus (Figure 5)
reached after
biomanipulation
carried out since
1990 (Table 2, no.
25, Figure 5)
xScheffer et al., 1992;
Noordhuis et al., 2016
19 Veluwe-meer 3400 5/1.55 1982-89: Flushing with
nutrient-poor water
1975-76
P. pectinatus
0-5%
(Figure 4)
1977-1995
P. pectinatus, P. perfoliatus,
P. pusillus, Characeae
10-15% (Figure 4)
1996-now
P. pectinatus,
C. aspera,
C. contraria,
N. obtusa
30-85% (Figure 4)
xScheffer et al., 1992;
Van den Berg et al.,
1998, 1999; Noordhuis
et al., 2016
20 Eemmeer 1520 ?/2.1 Since 1995: improved
sewage treatment and
closure of treatment plant
1970-1999 P. pectinatus
1–5% (Figure 4)
2000-ongoing P. crispus,
P. pectinatus, P. pusillus 10-40%
(Figure 4)
not yet reached,
but first Characeae
visible since 2010
(Figure 4)
xNoordhuis et al., 2016
21 Arresø DK 3987 5.6/3.1 Improved sewage
treatment, artificial lakes on
the main inlet stream,
reduced
catchment.fertilization
1989-1996
none
?- 2011: C. globularis, C. vulgaris,
M. spicatum, P. perfoliatus,
P. berchtoldii, P. cripus, P. pectinatus
unclear whether
already reached
xJeppesen et al.,
2007a,b Søndergaard,
unpubl. data
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Hilt et al. Response of Macrophytes to Restoration
TABLE 2 | The response of north temperate shallow lakes in The Netherlands (NL), Sweden (SE), Denmark (DK), United Kingdom (UK) ,and Germany (DE) to biomanipulation, natural fish kills or other lake-internal
measures (x: data on Secchi disk transparency and total phosphorus concentrations were available and used in Figure 2B).
No. Lake Country Size
(ha)
Depth
(max/mean)
(m)
Measures applied Period of turbid
conditions and major
remaining macrophyte
species
Periods of clear conditions,
macrophyte species and
coverage (% lake area)
Period of return
to turbid
conditions
Data
Figure 2B
References
22 Duiniger-meer NL 30 ?/1 1992,1993,1994: Fish
removal
?-1992 1992-ongoing
Chara globularis, C. vulgaris
Nitellopsis obtusa (2000-04)
Varying
macrophyte cover,
but no shift back
to turbid
xVan Berkum et al.,
1995; Meijer et al.,
1999; Riegman, 2007;
Verhofstad et al., 2017;
Van Donk unpubl. data
23 Ijzeren Man 11 2.3/2.3 1989: Complete removal of
fish biomass by pumping
dry, restocking with pike
fingerlings, roach, rudd, ide
and tench, sediments
removed
1960s−1989 Within 2 months of fish removal,
macrophytes covered 50% of lake,
Characeae
1995-?,
varying
macrophyte cover
xMeijer et al., 1999; Van
Donk unpubl. data
24 Noorddiep 4.5 ?/1.5 1988: Biomanipulation ?-1989 1989-?
E. canadensis,
C. demersum
clear for at least 8
years despite TP
250 µg L−1
xMeijer et al., 1999; Van
Donk unpubl. data
25 Wolderwijd 2650 5/1.5 1990: Biomanipulation See Table 1 (lake no. 18) 1992-
C. contraria, C. vulgaris
Figure 5
none x Meijer and Hosper,
1997; Noordhuis et al.,
2016
26 Zwemlust 1.5 2.5/1.5 1987: Lake drained empty,
fish completely removed,
restocked with pike and
rudd
1999: Temporary lowering
of water level, fish removal
?-1987 1988–1996
1988–1989: E. nuttallii,
1990–1991: C. demersum
1992–1994: P. berchtoldii
1995–1996: E. Nuttallii
1997–1999 x Van de Bund and Van
Donk, 2002;
Verhofstad et al., 2017
27 Terra Nova 85 ?/1.4 2003: (Removal of
planktivorous and
benthivorous fish)
1987–2003
1994: sparse stands of
C. demersum, P. lucens,
P. obtusifolius, P. pectinatus,
M. spicatum
2004-?
C. demersum, E. nuttallii,
P. obtusifolius, Nitella mucronata,
N. marina, Utricularia vulgaris
Van de Haterd and Ter
Heerdt, 2007
28 Galgje 3.1 ?/1.1 1987: Removal of all
planktivorous and 85% of
benthivorous fish in 1987
?-1988 1988: Within 2 months of fish removal
macrophyte covered lake, Characeae
2014: C. demersum, L. minor,
P. crispus, S. polyrhiza
Meijer et al., 1999;
Verhofstad et al., 2017
29 Loender-
veense
Plas
270 ?/2.7 2004/05: Removal of 95%
of fish stock
1980s−2004
P. perfoliatus
2005-present
E. nuttallii, N. marina, C. globularis,
C. connivens
none Pot and Ter Heerdt,
2014; Verhofstad et al.,
2017
30 Naarder-meer 1,042 ?/1.0 1993–1996: Sediment
dredging
1980–1989
Z. palustris, P. pectinatus
1990–1995-?
Najas marina, M. spicatum,
C. globularis, R. circinatus,
C. demersum, N. obtusa
Bootsma et al., 1999
(Continued)
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Hilt et al. Response of Macrophytes to Restoration
TABLE 2 | Continued
No. Lake Country Size
(ha)
Depth
(max/mean)
(m)
Measures applied Period of turbid
conditions and major
remaining macrophyte
species
Periods of clear conditions,
macrophyte species and
coverage (% lake area)
Period of return
to turbid
conditions
Data
Figure 2B
References
31 Finjasjön SE 1,100 12/2.7 since 1970: nutrient load
reduction, 1987:
suction-dredging of
sediments, 1992-2014.
removal of cyprinids
(Abramis brama, Rutilus
rutilus)
?-1994
M. spicatum, E. canadensis,
P. perfoliatus
1995-?
E. canadensis (disappeared 1997),
M. spicatum
P. perfoliatus
?Annadotter et al., 1999;
Strand and Weisner,
2001; Lage et al., 2015
32 Ringsjön
(Western Bay)
1,480 5.4/3.1 1992: Removal of about
50% of cyprinid fish
?-1992 P. crispus, P. perfoliatus,
P. pectinatus,
M. spicatum,
P. lucens, E. canadensis
2000–2005 Strand, 1999; Hansson
unpubl. data
33 Vasatorp-
dammen
2.1 1.4/1.1 1992: Fish removal by
rotenone
1989–1992 none 1993–1996
P. natans, P. obtusifolius,
C. demersum, C. globularis
?Blindow et al., 2000
34 Væng DK 15 1.9/1.2 1986-88 and 2007-09: Fish
removal
?-1986 1989–1996 and
2010-now
E. canadensis
1997–2009 x Jeppesen et al., 1991;
Søndergaard et al.,
2017
35 Arreskov 317 3.7/1.9 1991: Fish removal ?-1991 1992–1998
Characeae,
P. pectinatus, P. pusillus, P. crispus,
Z. palustris,C. demersum,
Z. pedunculata
xLauridsen et al., 2003b;
Søndergaard et al.,
unpubl. data
36 Alderfen UK 5.2 1.2/1 1979: Isolation from inflow,
1990: natural fish kill,
1992–1993: sediment
removal, 1995, 2000: fish
removal
Several turbid periods C. demersum
20–65%
several phases of
macrophyte
decline (e.g.,
1994, 1999–2000)
xMoss et al., 1986,
1990; Perrow et al.,
1997; Hoare et al.,
2008; Phillips et al.,
2015
37 Cockshoot
Broad
5.5 1.2/1 1992: Isolation from river,
sediment removal;
1989/90, 1996-2002,
2004-08:
fish removal
1970–1980
none
1990-2012
C. demersum, N. marina,
Z. palustris
40-58%
none x Moss et al., 1986;
Hoare et al., 2008;
Phillips et al., 2015
38 Hoveton Little
Broad Pound
End
15.5 1.5/1.0 1990: Suction dredging,
1990-1999: several fish
removals from isolated bay
(Pound End)
1970s−1991 1995-2006
C. demersum, N. marina
Low macrophyte
abundance since
2001
xHoare et al., 2008;
Phillips et al., 2015
39 Ormesby
Great Broad
40 1.5/0.9 1995: Fish removal 1970–1989
Chara, Z. palustris,
P. pectinatus, C. demersum,
M. spicatum, P. pusillus,
P. crispus
1995–2010
C demersum, E. canadensis,
Z. palustris, P. pectinatus, Chara,
P. friesii
67%
none x Phillips et al., 2015
(Continued)
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Hilt et al. Response of Macrophytes to Restoration
TABLE 2 | Continued
No. Lake Country Size
(ha)
Depth
(max/mean)
(m)
Measures applied Period of turbid
conditions and major
remaining macrophyte
species
Periods of clear conditions,
macrophyte species and
coverage (% lake area)
Period of return
to turbid
conditions
Data
Figure 2B
References
40 Cromes 4.3 1.2/1 1988: Sediment removal,
1992: Barrier to isolate from
river,1999: natural fish kill,
2004: Sediment removal
? 1995-ongoing
C. demersum
80%
xPerrow et al., 1997;
Phillips et al., 2015
41 Barton Broad 75 2/1 1980: P reduction from
effluents upstream 1996:
sediment removal
see Table 1 (lake no. 17) 2000-2012
C. demersum, P. pectinatus,
E. canadensis
12%
xPhillips et al., 2005,
2015
42 Schollener
See
DE 95 1/<1 2002: Natural fish kill during
summer flood
∼1980 to 2003 none 2004
C. demersum, N. marina,
P. berchtoldii, P. crispus
2005-ongoing? x Knösche, 2008
43 Rangsdorfer
See
272 2.5/1.5 2009/10: Natural winter fish
kill
?-2009
(probably several decades)
2010–2011
C. demersum, M. spicatum,
P. crispus
2012- ongoing Hussner et al., 2014
44 Schwandter
See
16.5 2.5/1.6 2002: P-precipitation with
aluminum sulfate
2009/10: Natural winter fish
kill
?-1995-2002
sparse stands of
C. demersum, P. crispus
2003-2010
C. demersum, P. crsipus
100%
2011–2015-?:
shift to turbid state
after carp stocking
xMathes, 2007; Nixdorf
et al., 2013; Hussner
et al., 2014
5 Ivenacker See 73.3 1.9/1.1 2009/10: Winter fish kill
2012/13: sediment dredging
?-2009 2010-?
Characeae
? x Nixdorf et al., 2013
46 Schloßsee
Buggen-
hagen
9.8 2.9/0.8 1990-96:
Improved wastewater
treatment,
1997: Sediment dredging
?-1997 1998-?, 2006-?
C. demersum, C. submersum.
C. hispida
2003 x Mathes, 2007
47 Möllener See 18.7 2.2/2 2006: P-precipitation with
aluminum sulfate
? 2007–2009-?
Characeae (occurred 3 years after
treatment)
?Hussner et al., 2014
48 Bachtel-
weiher
4.8 2/1.6 2002: Lake drained empty,
fish completely removed,
restocked with pike
?-2002 2003–2006-?
M. spicatum, R. trichophyllus,
C. globularis, N. mucronata
50-90%
?Hussner et al., 2014
49 Herren-wieser
Weiher
6.7 4.7/1.8 2001: Lake drained empty,
partial sediment removal,
fish completely removed,
restocked with pikeperch
?-2001 2002
Mainly R. circinatus, P. berchtoldii,
P. crispus, P. lucens, P. pectinatus,
C. contraria, C. globularis
75%
2003–2005-?
(after illegal carp
stocking)
Hussner et al., 2014
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Hilt et al. Response of Macrophytes to Restoration
FIGURE 2 | Total phosphorus (TP) concentrations and Secchi depth in spring
(April–June) and summer (July–September) of different north temperate
shallow lakes (A) before and after external nutrient load reductions during the
turbid, the intermediate recovery and the clear-water state (for details see
Table 1) and (B) before (turbid) and after (clear) biomanipulation or other
lake-internal measures (for details see Table 2).
restored through biomanipulation of the fish community
(Noorddiep, Wolderwijd, Zwemlust, for details see no. 24–26 in
Table 2).
Mann-Whitney U-tests were performed to compare lake
size, maximum and mean depths between lakes with different
restoration measures (external nutrient load reduction vs.
internal measures). Total phosphorus concentrations and Secchi
disk transparency were compared between the different states
(turbid, intermediate, clear) in lakes with external nutrient load
reduction) using Kruskal-Wallis tests and subsequent posthoc
comparisons (separately for different seasons). The same was
done for lakes with internal measures using Mann-Whitney U
tests. All statistical tests were run in SPSS.
PCLake Simulations
Simulating the response of water clarity and macrophyte biomass
to external nutrient load reduction and detecting thresholds of
nutrient loading for either intermediate recovery or clear state
required adaptation of the established ecosystem model PCLake.
This model has previously been used to estimate threshold
responses of shallow lakes to nutrient loading (Janse et al.,
2008; Janssen et al., 2017), and to simulate the response of
temperate shallow lakes to climate warming (Mooij et al., 2007),
to mowing of macrophytes (Kuiper et al., 2017) and—in a variant
of the model with three plant species—to biomanipulation and
herbivory (Janse et al., 1998). PCLake consists of a number
of coupled ordinary differential equations that describe the
most important biotic (submerged macrophytes, phytoplankton,
detritivorous macrozoobenthos, zooplankton, zooplanktivorous
fish, benthivorous fish, and piscivorous fish) and abiotic (detritus,
inorganic material, dissolved phosphorus, ammonium, and
nitrate) components of both the water column and the top-
layer of the sediment in a non-stratifying shallow lake (Janse,
1997, 2005). All organic components (apart from predatory
fish) are modeled in terms of dry weight (DW), nitrogen
(N), and phosphorus (P), and hence the nutrient-to-dry-weight
ratios of the organic components are variable. Internal fluxes
of nutrients between the sediment layer and the pelagic zone,
including internal loading, are accounted for and modeled
dynamically.
For our simulations we used the default settings of a lake in
PCLake. This default lake represents a relatively shallow lake with
an average depth of 2 m and is relatively small with a maximum
fetch of 1,000 m, an areal hydraulic loading of 20 mm days−1
(=7.2 m year−1), no infiltration or seepage, no surrounding
wetland zone, and a lightly clayish sediment (30% dry matter, of
which 10% organic matter, and 10% fine mineral material) (Janse
2005). Due to small size and shallowness, the lake is dominated
by macrophytes when nutrient loads are sufficiently low, but as
nutrient loads increase the lake switches to a turbid state. This
switch occurs rather suddenly due to the positive feedbacks in the
model (Janse 2005) that lead to a critical transition (e.g., Scheffer
and Carpenter, 2003). A common method to determine critical
transitions is bifurcation analysis. In this approach the model is
run to equilibrium several times, each with a different nutrient
load. For each run, the yearly average phytoplankton chlorophyll-
a concentration and macrophyte biomass are calculated. To
assess the presence of hysteresis this procedure is repeated twice
for each level of nutrient load, the first starting from a clear
lake and the second from a turbid lake. Where the equilibrium
outcomes of these two runs with identical nutrient load differ,
hysteresis is inferred. Here we ran the model for nutrient loads
ranging from 0.1 to 2.5 mg m−2days−1(0.4–9.0 kg ha−1year−1)
to cover a wide range of the eutrophication axis. The output of
the bifurcation analysis is a load-response curve (or bifurcation
plot) showing the effect of nutrient load on the biomass of
primary producers. The point of a sudden switch marks the
critical transition(s).
To simulate the influence of temperature and light on
the response of different macrophyte species to nutrient load
reduction we had to make two adjustments to the original
formulations of PCLake, while maintaining the modeling of
macrophytes as one functional group. First, the original power
function for temperature limitation of macrophytes was replaced
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Hilt et al. Response of Macrophytes to Restoration
by a temperature optimum curve LT(-):
LT=e−0.5
σ2h(T−Topt)2−(Tref −Topt )i
Here, σ(◦C) is the temperature constant based on a Gaussian
curve, T(◦C) is the water temperature, Topt (◦C) is the optimum
temperature for macrophytes and Tref (◦C) is the reference
temperature used to normalize the limitation function to 1
(Janse, 2005). With this function the model is flexible to simulate
macrophytes with different temperature optima. The second
adjustment is the timing of root allocation which occurs in
the autumn when macrophytes store energy to overwinter, for
instance, in propagules. In the original PCLake, this timing was
linked to a specific day in the year while submerged plants
are known to respond to physiological and environmental cues,
such as light availability, to determine timing of root allocation
(Madsen, 1991). Hence, we decided to link the timing of root
allocation to a minimum daily light availability for macrophytes,
following Madsen (1991),Van Dijk and Van Vierssen (1991), and
Van Dijk and Janse (1993). The available light for macrophytes is
based on the light availability over the depth of the water column
corrected for periphyton shading which is estimated by:
PAR =PAR0∗D−
LN PAR0
PARbot 
D∗1−εperip hyton
In which PAR (W m−2) is an approximation of the
average Photosynthetic Available Radiation (PAR) for plant
photosynthesis, PAR0(W m−2) is the PAR available at the top of
the macrophyte layer, PARbot (W m−2) is the PAR available at the
bottom of the macrophyte layer, D (m) is the depth and εperiphyton
is the shading by periphyton. In order to restrict complexity,
we refrained from adding periphyton as an extra compartment
to the model, but instead used the empirical relationship of
Vadeboncoeur et al. (2006) to estimate periphyton chlorophyll-a
biomass:
LOG10 Chlaperip hyton=c1∗LOG10 (TP)+c2
where c1=1.79 and c2=0.85 and TP is the in-lake
total phosphorus concentration (mg m−3). The periphyton
chlorophyll-a biomass is then used to estimate the shading effect:
εperi phyton =Chlaperip hyton ∗S∗f(L)∗f(T)∗f(α)
where Sis the specific light attenuation by periphyton (m g−1)
and correction factors for light limitation f(L), temperature
limitation f(T) and available plant surface area for periphyton
growth f(α). The shading effect of periphyton affects the timing
of root allocation through light availability to macrophytes as well
as the light limitation of macrophyte shoots.
PCLake has previously been calibrated following a Bayesian
approach to parameter estimation and uncertainty analysis using
data from nearly 40 temperate shallow lakes (Janse et al., 2010).
Although this calibration did not account for the specific effect
of periphyton, the data used for model calibration most likely
integrate this effect indirectly (Kuiper, 2016). By adding the effect
of periphyton to PCLake implicitly, we thus have to technically
recalibrate the model. This implies adjusting the parameter
settings of the model, such that given the same boundary
conditions, the model produces the same output. Therefore we
have calibrated the adjusted model manually by lowering the
half saturation light constant of vegetation and decreasing the
parameter for dark respiration of vegetation (see Table 3 for new
and calibrated parameter settings). PCLake is implemented in
DATM (Mooij et al., 2014). For the full overview of parameter
settings and model formulations, please see the DATM-file in the
Supplementary Material.
RESULTS
Lake Water Quality following External and
Internal Restoration
Our literature review provided information on water quality
and macrophyte development in 21 turbid lakes that were
subject to external nutrient loading reduction without
additional in-lake measures (Table 1) and 28 lakes with in-
lake restorative measures. Some of these measures were preceded
or accompanied by external nutrient loading reduction (Table 2).
Lakes with internal measures were on average smaller
than lakes with external nutrient load reduction alone, while
maximum depths were higher but mean depths were similar
(Table 4). Turbid conditions lasted from 1 year (Lake Veluwe)
to 51 years (Dümmer). Often, however, exact timing and
duration of the turbid period are unknown (Tables 1,2). During
the turbid phase, spring and summer TP concentrations were
high (∼>0.15 mg L−1), while Secchi disk transparencies
were low (∼0.4 m), with considerable differences between lakes
(Figures 2A,B,Table 4).
TABLE 3 | Parameter settings for PCLake.
Parameter Description Unit ValueaSource
σTemperature constant
based on a Gaussian curve
◦C 20 -
Topt Optimum temperature for
macrophytes
◦C 20 -
Tref Reference temperature ◦C 20 –
c1Slope of logistic curve
periphyton
– 1.79 Vadeboncoeur
et al., 2006
c2Intercept logistic curve
periphyton
–−0.85 Vadeboncoeur
et al., 2006
S Light attenuation by
periphyton
m g−10.03 Van Dijk, 1993
Lmin Minimal light availability cue
needed for plants to initiate
increased root allocation
W m−291.2 Calibrated
hLveg Half saturation light constant
of vegetation at 20◦C
W m−212 (17) Calibrated
kDresp Dark respiration rate of
vegetation
day−10.015 (0.02) Calibrated
aValues between brackets are the original value.
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Hilt et al. Response of Macrophytes to Restoration
TABLE 4 | Size, maximum and mean depths, total phosphorus concentrations and Secchi disk transparency in lakes after external nutrient load reduction or implementation of internal measures.
Lakes State Size (ha) Maximum depth (m) Mean depth (m) Total phosphorus concentrations (µg L−1) Secchi disk transparency (m)
spring summer spring summer
mean median mean median mean median mean median mean median mean median mean median
External nutrient
load reduction
Turbid 984 450 5.0 3.9 2.4 1.8 281 163 A 317 244A 0.66 0.45 A 0.46 0.4 A
Intermediate 97 94 A 206 206 A 1.23 1.11 AB 0.75 0.6 AB
Clear 53 51 B 59 49 B 1.74 2.17 B 1.17 1.12 B
P(Kruskal-Wallis test) 0.02 0.007 0.042 0.02
Internal measures Turbid 273 18 2.8 2.1 1.5 1.3 229 156 a 256 150 a 0.62 0.45 a 0.45 0.40 a
Clear 168 81 a 170 103 b 1.12 0.96 b 1.10 0.99 b
P(Mann-Whitney U test) <0.001 0.016 0.076 0.051 0.011 0.026 <0.001
Size and depths include all lakes, while TP and transparency data were only available for selected lakes (see Figure 2,Tables 1,2). Different letters indicate significant differences between values of different states (external nutrient load
reduction: Kruskal-Wallis test, capital letters; internal measures: Mann-Whitney U-test, small letters).
External nutrient loads were usually reduced following
improved sewage treatment in the catchment. Lake Veluwe and
Wolderwijd were flushed with nutrient-poor water (Table 1). The
most commonly applied internal measure was biomanipulation
in the form of removal of benthivorous and planktivorous fish
(in some cases by draining/pumping the lake dry). In nine
lakes, sediment removal was applied as an additional or the sole
measure and in five lakes natural fish kills occurred during severe
winters or as a result of a summer flood (Table 2), emulating
the effects of planned biomanipulation. During the first years
after nutrient load reduction, TP concentrations were lower than
during the turbid period, but still about twice as high in summer
as in spring. Spring water transparency was higher than during
the turbid period, while summer values were still low (Table 4,
Figure 2A). This phase has been found to last up to 20 years
(e.g., Müggelsee), but often its start has not been recorded and/or
lakes have not yet reached stable clear conditions (Table 1).
Lakes Müggelsee, Veluwemeer, and Eemmeer show a similar
intermediate recovery state with spring water transparencies
being higher than during turbid summer conditions which lasted
for about 20 years (Figure 3). A switch back from intermediate
to turbid conditions has only been observed in Lake Steinhuder
Meer, in this case about 10 years after macrophytes returned
(Table 1).
After implementation of in-lake measures, TP concentrations
were at the same level as before, in both spring and
summer, whereas water transparency in spring and summer was
significantly higher than before restoration (Table 4,Figure 2B).
These unstable clear conditions often only lasted for a few
years and many lakes shifted back to turbid conditions (e.g.,
no. 23, 26, 32, 34, 36, 38, 42, 43, 44, 46, 49 in Table 2). Fish
stock reductions in Wolderwijd, Zwemlust and Noorddiep led
to clear conditions both in spring and summer (Figure 4). Lake
Zwemlust shifted back to turbid conditions after 9 years, while
Noorddiep stayed clear for at least 8 years with no further
information on subsequent periods.
Only six of the lakes with external nutrient load reduction
(∼25%, no. 1, 3, 7, 10, 18, and 19 in Table 1) reached stable
clear-water conditions with lower ambient TP concentrations
and higher Secchi depths than during the intermediate recovery
state both in spring and summer (Figure 2A,Table 4). In
four of the 28 lakes (14%, no. 25, 29, 37, 39 in Table 2,
Lake Wolderwijd in Figure 4), stable, longer-term clear-water
conditions with a diverse macrophyte flora were obtained after
the application of in-lake measures. For several lakes, their
longer-term development is not known (Table 2).
Model Simulations on Lake Response to
External Nutrient Load Reduction
The results of the simulations using the adjusted PCLake model
revealed three stages for lakes that undergo external nutrient
load reduction: a turbid state, an intermediate recovery state and
a clear state (Figure 5A). For the default lake in PCLake, the
turbid state occurs if the P load exceeds 1.3 mg P m−2days−1.
The intermediate recovery state occurs between the two critical
transitions that appear at a P load of 1.06 and 1.3 mg P m−2
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Hilt et al. Response of Macrophytes to Restoration
FIGURE 3 | Total phosphorus (TP) concentrations and Secchi depth in spring (April–June) and summer (July–September) and macrophyte coverage in Lake
Müggelsee; Lake Veluwe and Lake Eem during the turbid (green), the intermediate recovery (brown) and the clear-water (blue) state (for lake details see Table 1).
days−1. If the P load drops below 1.06 mg P m−2days−1the lake
turns into a clear state. The critical P loads for the intermediate
recovery state are smaller in the adjusted model including
periphyton than with the original formulation of PCLake (critical
transition between approximately 1 and 2 mg P m−2days−1,
Janse et al., 2008). The periphyton effect thus reduces the
threshold for intermediate recovery state of the default lake by
about a quarter of what it would be without periphyton. The
reason for this is that in case of hysteresis, the position of the
highest critical nutrient load is mainly determined by macrophyte
characteristics, while the position of the lowest critical nutrient
load is mainly determined by phytoplankton characteristics. This
is in agreement with the results of the sensitivity analysis of
PCLake (Janse et al., 2010). Since periphyton negatively affects
the performance of macrophytes, the highest critical nutrient
load was reduced, while phytoplankton characteristics were
unaltered and the lowest critical nutrient load (1 mg P m−2
days−1) thus did not change.
The turbid state was characterized by lack of macrophytes
and enhanced phytoplankton biomass with increasing nutrient
loading (Figure 5A). The intermediate state was characterized by
a changed phenology of the primary producers. With increasing
nutrient loading, the phytoplankton summer peak shifted to
earlier dates of the year and this advancement of phytoplankton
was mirrored by an abbreviated macrophyte growing season
(Figure 5A). The shorter growing season was a direct effect
of the inclusion of periphyton shading in our adapted version
of the model. The shading of macrophytes by periphyton was
most severe at the peak of summer when the light input and
water temperature are high. As a result, macrophyte growth was
limited to the period just after the clear water phase in spring
until the start of the summer phytoplankton bloom. The clear
state was characterized by an increase of macrophyte biomass
with increasing nutrient loading (Figure 5A). Phytoplankton
production was restricted to the spring bloom peak and there was
no summer bloom.
The bifurcation plot (Figure 5B) shows a less sudden
transition of macrophytes and phytoplankton compared to the
abrupt transition between the phytoplankton-dominated turbid
state and the macrophyte-dominated clear state, often seen in
bifurcation plots in literature (e.g., Janse et al., 2008). The
gradual course of the bifurcation plot is due to the inclusion
of periphyton shading of macrophytes. This shading permits
high biomass of both macrophytes and phytoplankton within
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Hilt et al. Response of Macrophytes to Restoration
FIGURE 4 | Total phosphorus (TP) concentrations and Secchi depth in spring (April–June) and summer (July–September) and macrophyte coverage in Lake
Wolderwijd, Lake Zwemlust and Lake Noorddiep before (green and brown) and after (blue) biomanipulation (for lake details see Table 2). In Lake Zwemlust,
P. berchtoldii occurred instead of P. pectinatus, and the coverage in Lake Noorddiep was only estimated based on the information that it was higher than 25% (Gulati
and Van Donk, 2002).
the same year during the crashing or intermediate recovery
phase. Furthermore, the region of hysteresis is tilted, leading
to a less abrupt and thus more realistic critical transition from
phytoplankton dominance to macrophyte dominance and back
again.
Macrophyte Species Recovery following
External and Internal Restoration
During the turbid phase, three lakes with subsequent external
nutrient load reductions were reported to lack macrophyte
stands altogether and six and one lakes had sparse stands of
Potamogeton pectinatus (also known as Stuckenia pectinata) and
P. pusillus, respectively, while no information was available
for the remaining lakes (Table 1). For lakes with internal
measures, information on macrophyte species present during the
turbid phase is available for 10 lakes. Apart from P. pectinatus
and P. perfoliatus, plants such as Ceratophyllum demersum,
M. spicatum or E. canadensis are mentioned, if indeed plant
stands were present at all (Table 2).
During the intermediate recovery state P. pectinatus was the
dominant macrophyte species in two thirds of the analyzed lakes
with reduced external nutrient loading. Other pondweed species
such as P. perfoliatus, P. crispus and P. pusillus or Zannichellia
palustris were also found in several lakes during the intermediate
recovery state, while other groups such as Characeae or Elodea
species were much less common (Table 1,Figure 6). Lakes
Müggelsee, Veluwemeer and Eemmeer were all dominated by
P. pectinatus during the intermediate recovery state which lasted
for about 20 years (Figure 3). Müggelsee and Veluwemeer seem
to have entered a stable clear state with more diverse submerged
vegetation (including Characeae in Veluwemeer) in 2011 and
1996, respectively, while Eemmeer has not yet reached that
phase despite the recent detection of Characeae (Figure 3). Three
other lakes also reached a stable clear state and were colonized
by different species of Characeae and/or Najas marina, Elodea
species and C. demersum (Table 1).
The dominant macrophytes occurring after lake internal
measures were Characeae, C. demersum, Elodea species or
N. marina (Table 2,Figure 6). Often, lakes had only one or two
dominant species, and in at least 10 cases, lakes switched back
to turbid conditions and lost their macrophytes again (Table 2,
Figure 4). The response of macrophytes occurred gradually in
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Hilt et al. Response of Macrophytes to Restoration
FIGURE 5 | (A) Time series of simulation using PCLake of chlorophyll-aconcentrations and macrophyte shoot biomass for different phosphorus (P) loadings within
the clear, intermediate recovery and turbid states. Darker colors are associated with higher P loading simulations. (B) Hysteresis plots showing yearly mean simulated
values of chlorophyll-aconcentrations and macrophyte shoot biomass for different phosphorus (P) loadings within the clear, intermediate recovery and turbid states.
Arrows denote the directions of the hysteresis effects.
Wolderwijd, showing an increase in Characeae coverage, while
in the much smaller Zwemlust and Noorddiep, macrophytes
immediately covered large areas, E. nuttallii, E. canadensis and
C. demersum dominating (Figure 4,Table 2). In four lakes, stable
clear-water conditions were obtained after in-lake measures, all
being characterized by Characeae dominance (Table 2,Figure 4).
DISCUSSION
Our analysis of long-term data from 49 temperate shallow
lakes during their recovery from a turbid phase reveals
that both a reduction of the external nutrient loading and
implementation of lake-internal measures often result in the
occurrence of an intermediate state (Figure 1) that can last
for several decades. External nutrient load reductions are often
followed by the re-occurrence of a spring clear-water phase
that opens a “window of opportunity” for macrophyte re-
colonization, but with only a short growth period due to turbid
conditions during summer. This pattern was confirmed by our
model simulations (Figure 5). As hypothesized, macrophyte re-
establishment following nutrient load reduction occurs in a
reversed sequence to the one described for eutrophication by
Sayer et al. (2010a,b). In contrast, lake internal measures such as
fish or sediment removal often result in clear-water conditions
during spring and summer. This clear-water state is, however,
often only temporary and lakes frequently shift back to turbid
conditions one or a few years after the restoration. Likely, the
duration of the clear-water conditions is related to nutrient
loading and the intensity of the restoration effort (Hansson et al.,
1998). Only in a few examples have longer lasting clear-water
conditions been observed. These required spring and summer TP
concentrations below 0.05 mg L−1.
We also have evidence for our second hypothesis, namely
that different types of restoration measures influence the
macrophyte community composition. P. pectinatus and a few
other pondweeds most often recolonise temperate shallow
lakes with reduced external nutrient loading and dominate
during the intermediate recovery state. The implementation of
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Hilt et al. Response of Macrophytes to Restoration
FIGURE 6 | The 10 most common macrophyte taxa in north temperate
shallow lakes during re-colonization after reduction of external nutrient loading
(light gray) or implementation of internal measures (dark gray; for more details
see Tables 1,2).
internal restoration measures results in the establishment of
a different community, often consisting of a small selection
from either hornwort (C. demersum), charophytes, water weeds
(E. canadensis, E. nuttallii) or naiad (N. marina). Only in a
few cases have lakes reached a state of clear-water conditions
during spring and summer and with a more diverse macrophyte
community.
Submerged Macrophyte Survival during
the Turbid Phase
Whether and which macrophyte stands or propagules survive
during the turbid phase depends on the occurrence of
macrophyte species before the shift to turbid conditions and on
the length and severity of the turbid phase (e.g., Vari and Toth
2017). Seed banks in shallow lake sediments have often been
assumed to be insufficient for recovery of submerged vegetation
by germination, due to low numbers of viable seedlings, strong
seed dormancy, strict germination cues and the reliance of many
species upon vegetative reproduction (Haag, 1983; Kautsky, 1990;
Rodrigo et al., 2013; Baldridge and Lodge, 2014). In contrast,
De Winton et al. (2000) and Verhofstad et al. (2017) have
shown that seed banks from even the most degraded lakes are
capable of an emergence response and thus offer a potential
means to restore vegetation. In our survey, the duration of
complete macrophyte loss was often unrecorded, but periods of
several decades are common (Tables 1,2). If macrophyte stands
survived during this period and were recorded, these were often
sparse stands of P. pectinatus, a species commonly associated
with the crashing phase during eutrophication (Sayer et al.,
2010a). This species survives in very shallow water even under
phytoplankton dominance (Hilt et al., 2013) and its strongly
apical growth form may allow it to survive at greater depths
in turbid water. The shallow littoral, especially in larger lakes,
is strongly disturbed by wave action and only species with
high anchorage and breaking strength can survive under these
conditions (Schutten et al., 2005). Potamogeton pectinatus has
a high breaking strength (Brewer and Parker, 1990) and its
phenotypic plasticity allows it to form short plants in shallower
water (Idestam-Almquist and Kautsky, 1995). In contrast, other
common species in eutrophic lakes such as E. canadensis or
M. spicatum have been described as deep water species with lower
tensile strength (Brewer and Parker, 1990), while C. demersum
has no roots for anchorage. These species are thus less likely
to persist through severely turbid states in very shallow littoral
areas and to serve as remnant populations for re-colonization, at
least in larger lakes. Nevertheless, they have been reported during
turbid phases in four lakes included in our survey, most probably
in wind-protected areas or bays.
Knowledge of the survival of propagules in sediments during
turbid phases is limited. In general, charophyte oospores and
macrophyte seeds have been found to survive up to 150 years
(Kaplan and Muer, 1990; De Winton et al., 2000; Alderton et al.,
2017). Germination tests with sediments have been suggested
before implementing lake restoration measures to forecast the
potential for macrophyte recovery from internal sources (Hilt
et al., 2006), however, these are not routinely applied. Re-
colonizing macrophyte clones of several species that originated
from periods before eutrophication have been found (Sand-
Jensen et al., 2008), but knowledge about the origin of re-
colonizing macrophytes post-restoration remains scarce (Bakker
et al., 2013).
Response of Macrophytes to Nutrient Load
Reductions
Shallow temperate lakes often show a relatively rapid response
to reductions in external phosphorus loading, characterized by
a reduction in phytoplankton biomass during spring and early
summer (Jeppesen et al., 2005). During late summer, however,
the response is delayed because of sustained remobilisation of
phosphorus from the sediment (Søndergaard et al., 2013). As a
consequence, high phytoplankton and cyanobacterial abundance
are often reasserted in summer (Sommer et al., 2012) resulting in
turbid water and preventing macrophyte growth. In our survey,
such conditions occurred at spring TP concentrations of around
0.1 mg L−1, while summer concentrations were twice as high.
The increased water transparency in spring and early summer
seems to be exploited by certain macrophyte species, in our
survey mainly P. pectinatus along with P. perfoliatus, P. crispus
and Z. palustris. These macrophyte species are characterized by
specific traits that may explain their prevalence. Firstly, they can
compress their whole life cycle into the short clear-water period
in spring and early summer due to early germination from tubers,
turions or seeds, shortening time to peak biomass and allowing
early formation of overwintering tubers and seeds (Table 5).
Secondly, they can have short growth forms that can establish
in very shallow habitats (e.g., Van Vierssen, 1982a). Thus, they
are often the species that survive in turbid conditions in shallow
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Hilt et al. Response of Macrophytes to Restoration
TABLE 5 | Characteristics of macrophyte species/groups typically recolonising temperate shallow lakes after external nutrient load reduction or following implementation of lake-internal measures such as
biomanipulation.
Restoration
measure
Macrophyte
species
Re-production
mode
Germination Colonization
mode
Timing of peak
biomass
Specific features Epiphyton
density
Allelopathic
activity
Susceptibility to
herbivory
External nutrient
load reduction
Potamogeton
pectinatus
Mainly by tubers,
although seeds
are formed1
Between
10-15◦C7
Rhizomatic
growth11
June19 Concentration of
biomass below
water surface1
Tubers rich in
carbohydrates32
Low13 Low2High3,14,28
P. perfoliatus Turions29, seeds Rhizomatic
growth11
High20
P. crispus Mainly by
turions26
Turions can sprout
in late summer, but
shoot elongation
after winter above
10◦C26
Turions Late spring to
early summer25
Low2
P. pusillus Turions4Rhizomatic
growth4,12
Low13 Low2High14
Zannichellia
palustris
Seeds5Mainly
temperature-
dependent, above
12-16◦C5,27
Seeds12 Short, but
numerous shoots
allow development
in very shallow
water27
Tolerant to
disturbance by
wave action27
Low13 Low2
Internal measures Elodea canadensis Vegetatively6May-June22, can
be evergreen
Fragments,
vegetative growth
(peripheral
propagation)4,12
Sudden biomass
collapses6,8,9
High13 Medium2Medium3
E. nuttallii Vegetatively17 Can be evergreen Fragments,
vegetative
growth12
Aug.-Oct.21 Extremely high
reproductive
capacity16,17
Sudden biomass
collapses
High13
Low15
Medium2Medium18
Ceratophyllum
demersum
Vegetatively
(dormant apices)
April22, can be
evergreen
Fragments,
vegetative
growth12
August31 Sudden biomass
collapses10
Low15 High2Low
Najas marina Annual, seeds Late in (temperate)
season (>20◦C)30
Seeds11 Mid August11 Medium2Low18
Characeae Oospores Can be evergreen,
overwintering by
shoot apices
Oospores Late summer23 Low15 Medium2Medium3
1Van Wijk (1989),2Hilt and Gross (2008),3Dorenbosch and Bakker (2011),4Barrat-Segretain and Bornette (2000),5Bytnerowicz and Carruthers (2014),6Simberloff and Gibbons (2004),7Madsen and Adams (1988),8Rørslett et al.
(1985),9Strand and Weisner (2001),10Van de Bund and Van Donk (2002),11Vari and Toth (2017),12Capers (2003),13Lalonde and Downing (1991),14 Hidding et al. (2010),15Grutters et al. (2017),16 Simpson (1990),17 Josefsson
(2011),18Pot and Ter Heerdt (2014),19Körner (2001),20 Choi et al. (2002),21Best and Dassen (1987),22 Best (1977),23Talling and Parker (2002),24Lillie et al. (1997),25 Woolf and Madsen (2003),26Tobiessen and Snow (1984),27Van
Vierssen (1982b),28Weisner et al. (1997),29 Wolfer and Straile (2004),30Van Vierssen (1982c),31Best (1982),32 Spencer (1986).
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Hilt et al. Response of Macrophytes to Restoration
margins (see Submerged Macrophyte Survival during the Turbid
Phase) and then expand into deeper water with improvements in
clarity during nutrient load reduction. Rhizomatic growth from
remaining P. pectinatus stands has been shown, via microsatellite
analyses, to be the dominant re-colonization mode in Lake
Müggelsee; more recently established P. pectinatus stands had
lower genotype diversity and were comprised of only a small
subset of genotypes from shallower areas (Hilt et al., 2013).
Thirdly, energy reserves in vegetative propagules such as tubers
of P. pectinatus allow early onset of growth independent of
light availability (Spencer, 1986). In addition, P. pectinatus can
concentrate large parts of its biomass just under the water surface
and thus survive in relatively turbid water (Van Wijk, 1988). An
initial colonization of formerly turbid lakes with P. pectinatus
during recovery has also been observed in deeper, stratifying
lakes. Thus, in Lake Tegel (Germany), this species dominated for
more than 20 years after the start of phosphorus stripping in the
major inflow (Hilt et al., 2010).
Usually, the maximum colonization depth of macrophytes
during the intermediate recovery phase is low (around 1 m)
and consequently, depending on the lake morphometry, only
small parts of the lake bed might be covered. In contrast,
very shallow lakes may reach over 50% cover (Table 1). This
suggests that in “deeper” shallow lakes, macrophyte coverage
during the intermediate phase may be insufficient to stabilize
clear-water conditions during later summer. Based on the
findings of Søndergaard et al. (2016) submerged macrophyte
coverage on average needs to pass a threshold of 20% of lake
area to markedly lower phytoplankton densities. In principle,
small stands can be sufficient as a refuge for phytoplankton-
grazing zooplankton against fish predation (Lauridsen et al.,
1996; Portielje and Van der Molen, 1999). However, abundant
colonial and filamentous cyanobacteria which often dominate
the summer phytoplankton communities during lake recovery
cannot be effectively controlled by zooplankton grazers (Wang
et al., 2010 and references therein). Bottom-up stabilizing
mechanisms of macrophytes on water clarity such as nutrient
competition, increased sedimentation within stands and reduced
sediment resuspension will be inefficient at low plant coverage
(Blindow et al., 2014). Low coverage is, however, not the
only reason why macrophytes in the intermediate recovery
phase cannot stabilize clear-water conditions in late summer,
as shown in the case of the very shallow Lake Dümmer, where
cyanobacteria blooms still occurred in summer despite high
macrophyte coverage.
Our model simulations suggest that high periphyton shading
triggers macrophyte disappearance in summer. Periphyton
shading, often accompanied by herbivory (Hidding et al., 2016),
has been shown to impair macrophyte development in empirical
studies (e.g., Jones et al., 2002; Jones and Sayer, 2003; Roberts
et al., 2003) and is argued to be a major factor in the failure
of macrophytes to establish even decades after the start of
nutrient loading reduction, despite suitable water clarity for
plant re-establishment in spring (Phillips et al., 2005). In our
adapted PCLake model, periphyton biomass was dependent on
TP concentrations in the water, based on the positive correlation
between chlorophyll content of periphyton on hard substrata
and TP in the water column (Vadeboncoeur et al., 2006). In
eutrophic shallow, temperate lakes, periphyton is often top-
down controlled by a cascading effect from omnivorous fish
that feed on periphyton grazers such as snails and chironomid
larvae (Jones and Sayer, 2003). Thus, nutrient load reductions
will only reduce periphyton shading after the fish biomass built
up during the turbid period has also been reduced, which
may take 10–15 years (Jeppesen et al., 2005). Furthermore,
the observed dominant macrophyte species in the intermediate
recovery phase after nutrient load reduction (Table 1) show little
or no allelopathic activity that might hamper periphyton growth
(Table 5), thus making them more susceptible to shading by
periphyton.
Cyanobacteria have been shown to potentially inhibit
submerged macrophyte growth via allelopathy (Zheng
et al., 2013), but whether this mechanism contributes to
the disappearance of macrophytes during the recovery phase in
summer is unknown. Most of the dominant macrophyte species
during intermediate recovery after nutrient load reductions are
also highly susceptible to herbivory due to their low content of
polyphenols, low carbon to nitrogen ratio and low dry matter
content (Elger and Willby, 2003; Dorenbosch and Bakker,
2011,Table 5). Periphyton shading may further increase the
sensitivity of macrophytes to herbivory (Hidding et al., 2016).
Finally, fine-leaved species such as P. pectinatus,P. pusillus and
Z. palustris also suffer from leaf plucking by omnivorous fish
during periods of low zooplankton abundance when those fish
switch to macroinvertebrate prey found in the periphyton of
macrophytes (Körner and Dugdale, 2003). Such leaf plucking by
fish can lead to a considerable leakage of nutrients from injured
macrophyte tissue, thereby further stimulating phytoplankton
growth (Hansson et al., 1987). Overall, while being well-suited
for survival during turbid phases and for exploiting the clear-
water conditions in spring for re-colonization, other traits of
macrophyte species typical of the intermediate recovery phase
following nutrient load reduction prevent their survival during
later summer (Table 5).
Stable clear-water conditions in spring and summer with more
diverse macrophyte vegetation were observed when both spring
and summer TP concentrations reached about 0.05 mg L−1. This
value corresponds well with a threshold for low cyanobacterial
abundance in shallow lakes found by Jeppesen et al. (2005) and
Triest et al. (2016) and the average critical loading for shifts
from turbid to clear conditions estimated for Dutch shallow
lakes by Janse (2005) and in eastern England by Phillips et al.
(2015). Whether external nutrient load reductions alone were
responsible for the observed low in-lake TP concentrations
in the lakes in our survey that reached stable clear-water
conditions, however, remains questionable. It seems that in
most cases additional changes in either the fish community
(Lake Veluwe, Galenbecker See) and/or exotic mussel invasions
(Lake Müggelsee, Eemmeer) contributed to the observed trend.
In Lake Veluwe, several severe winters, an increase in bream
(Abramis brama) fisheries between 1993 and 1997 and the
increase in zebra mussel densities are all thought to have
contributed to a break in the dominance of cyanobacteria, thus
allowing for the prevalence of stable clear-water conditions
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Hilt et al. Response of Macrophytes to Restoration
with charophyte dominance since 1996 (Noordhuis et al., 2016,
Figure 4). Once established, these dense charophyte beds provide
more efficient stabilizing mechanisms for clear-water conditions
than rooted angiosperms (Blindow et al., 2014). Characeae also
successfully replaced P. pectinatus in Lake Wolderwijd after
biomanipulation (Figure 4), while in Swedish Lake Krankesjön
a similar development has been observed, the reasons for which
are unknown (Blindow, 1992; Hargeby et al., 1994; Hansson et al.,
2010). In Lake Müggelsee and Eemmeer, the additional influence
of a sudden invasion of the quagga mussel (Dreissena rostriformis
bugensis) in around 2013 might have contributed to a decline in
TP concentrations and increased water transparencies (Figure 4,
S. Hilt, unpublished, Noordhuis et al., 2016). This species can
colonize soft substrates and thus cover much larger areas than
those previously occupied by the zebra mussel (D. polymorpha)
(Karatayev et al., 2015). In Lake Eemmeer, quagga mussels
filtered the lake volume about five times a day in 2013 (Noordhuis
et al., 2016).
Macrophyte recovery in Steinhuder Meer and Langer See
deviates from the suggested pattern in that both are dominated
by species more typical of lakes having undergone restoration
with internal measures (Table 1). In Steinhuder Meer, a
strong reduction of the fish population has been observed
which was attributed to cormorant activities (Niedersächsischer
Landesbetrieb für Wasserwirtschaft, Küsten- und Naturschutz,
2011). Cormorant effects on fish populations are also suggested
for Felbrigg Lake (C. Sayer, unpublished). Through a natural
increase in cormorants, the lake food web configuration
was likely affected in ways comparable to those of lakes
undergoing biomanipulation. This makes Steinhuder Meer and
Felbrigg Lake cases of external load reduction with added
unintentional internal measures (natural biomanipulation)
possibly accelerating recovery. Therefore, these cases show closer
correspondence to macrophyte community patterns of lakes
having undergone internal measures. They also illustrate that
parallel biological processes may be at play in recovering lakes
that need to be considered in unison to understand the speed and
trajectory of macrophyte recovery.
Response of Macrophytes to
Biomanipulation in Shallow Temperate
Lakes
In contrast to lakes undergoing only reduced external nutrient
loading, submerged macrophytes often respond very quickly
in shallow lakes subjected to biomanipulation by fish removal
(Hansson et al., 1998; Bakker et al., 2013), even at rather
high nutrient concentrations (Figure 2). Macrophytes colonizing
these lakes are often “pioneer” species, such as Elodea or
Ceratophyllum, re-colonizing from either seeds, oospores or
fragments and characterized by high growth rates (Tables 2,4).
Similar species have also been recorded in lakes following
natural fish kills (Sayer et al., 2016) or implementation of
other in-lake restoration measures such as sediment dredging
and phosphorus precipitation (Table 2). Fish removal may
indirectly (due to more periphyton grazing invertebrates) reduce
periphyton shading in summer, a major mechanism preventing
macrophyte survival after nutrient load reduction (see Response
of Macrophytes to Nutrient Load Reductions). The relevance
of this process for macrophyte recovery after fish removal
has not yet been directly tested, although mesocosm trials
in which the density of periphyton grazers are manipulated
produce predictable outcomes in terms of periphyton biomass
and macrophyte composition (Elger et al., 2009). Excretion of
allelopathic substances, which has been detected for many of the
typical species that colonize after biomanipulation (Table 5), may
also contribute to lower periphyton densities.
In general, macrophyte species typically occurring after
introduction of in-lake restoration measures allow for a longer
period of high macrophyte cover and dampen seasonal changes
in phytoplankton abundance as described for “stable” lakes prior
to major eutrophication (Sayer et al., 2010b). Both, Elodea and
charophyte species can remain evergreen in temperate lakes (e.g.,
Søndergaard et al., 2017), thus extending their positive influence
on water quality to seasons outside the influence of annual
species.
In many cases, however, mass developments of monocultures
occur. Monocultures of Elodea or Ceratophyllum species are often
unstable in terms of interannual persistence (Table 5) and can
collapse leading to a shift back to turbid conditions as in, for
instance, Lakes Zwemlust, Væng, and Alderfen Broad (Table 2).
Characeae seem less often involved in sudden collapses, although,
exceptions are known, for example Schlosssee Buggenhagen
(Table 2) or Lake Botshol (Rip et al., 2007). If lakes remain
clear for several consecutive years, which is usually only the case
at lower nutrient concentrations, a more diverse macrophyte
community develops (Table 2,Lauridsen et al., 2003a). Lauridsen
et al. (2003b) assumed that differences in the success of
biomanipulation in Danish and Dutch shallow lakes might be
attributable to variation in pioneer macrophyte species; thus,
Elodea and Potamogeton species, typical for Danish lakes, were
preferred over charophytes by macrophyte-grazing waterfowl
(Weisner et al., 1997). Indeed, increasing top-down control
of periphyton-grazing invertebrates by omnivorous fish, which
increase in abundance in the period after a biomanipulation,
may render macrophytes more susceptible to herbivory (Hidding
et al., 2016).
Conclusions and Implications for Lake
Management
Our analyses suggest that the composition of the macrophyte
community and their seasonal abundance in shallow lakes
during recovery from turbid, highly eutrophic conditions often
depends on remnant macrophyte stands, the specific restoration
measure applied and additional stochastic influences on water
clarity such as winter fish kills, cormorant predation on fish
or introduction of invasive filter-feeding mussel populations. In
turn, the prevailing macrophyte community can influence lake
water quality.
Reductions in external nutrient loading often result in
the re-occurrence of spring clear-water phases exploitable by
a few macrophyte species (mainly pondweeds) with specific
traits. Resistance to wave action permits survival during the
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Hilt et al. Response of Macrophytes to Restoration
turbid phase in very shallow areas, in particular in larger
lakes. During recovery these plants germinate early in spring
from energy-rich vegetative propagules and complete their
life cycle in early summer, when phytoplankton takes over.
This intermediate recovery phase may, in some cases, last for
several decades before a more diverse and abundant submerged
macrophyte community develops that stabilizes clear-water
conditions during the entire potential growing season (Figure 1,
Table 1). Our model simulations suggest that, if the premature
termination of macrophyte growth can be prevented, the summer
phytoplankton peak responsible for turbid water and potentially
harmful algae blooms will also be reduced. Simulations also
revealed that at high periphyton shading, the intermediate
recovery phase is shifted to lower nutrient loads compared with a
scenario with lower periphyton shading. Therefore, if periphyton
shading can be reduced external restoration measures could
potentially be effective at a higher nutrient load. Macrophyte
recovery during the intermediate recovery state might be
facilitated by establishing exclosures to protect certain areas from
herbivory by birds and/or predation of periphyton grazers by
omnivorous fish, a lake-wide biomanipulation of fish, or internal
measures, such as TP precipitation to lower water column
TP concentrations in summer. Additional, usually unintended
internal changes, such as reductions in fish abundance by
commercial fisheries, natural fish kills or exotic mussel invasions
can facilitate a shift to clearer conditions in summer and
further aid the establishment of a more diverse macrophyte
community.
In contrast, fish stock reductions, natural fish kills and
sediment removal via suction-dredging can, by themselves,
temporarily restore clear-water conditions in spring and summer,
even at high nutrient concentrations and then allow rapid
colonization by pioneer macrophyte species from in situ seeds,
oospores or vegetative fragments. Fish stock reductions might
thus be a suitable short-term management strategy, but fish
removal needs to be frequently repeated.
Lasting macrophyte recovery can only be achieved in
combination with reduced nutrient loading (Figure 1), a need
that is further accentuated under future climate change scenarios,
where cyanobacterial shading of macrophytes will likely be more
severe (Kosten et al., 2012). Although global, political measures
against ongoing climate warming are slow, local restoration
efforts may reduce the combined stress from e.g. eutrophication
and climate warming (Moss et al., 2011; Scheffer et al., 2015),
and thereby serve as a buffer against further deterioration of
macrophyte beds and the ecosystem services that derive from
lakes and reservoirs (Urrutia-Cordero et al., 2016).
AUTHOR CONTRIBUTIONS
SH conceived the presented idea, wrote the manuscript and
performed the literature research. MA, EB, IB, TD, L-AH, EJ, TK,
AK, JK, TL, RN, GP, JR, H-HS, MS, KvdW, EvD, AW, NW, and
CS provided lake data. MG, JJ, AJ, WM, and ST performed the
modeling. All authors contributed to discussions and the writing
of different parts of the text.
FUNDING
MA and MG were supported by the German Research
Foundation (DFG, grant no. SU 623/1-1 and GRK 2032/1,
respectively). AJ is supported by the Netherlands Environmental
Assessment Agency (PBL) and ST by STOWA (grant no.
443.269). L-AH was supported by the BiodivERsA ERA-
net LIMNOTIP. EJ, MS, TD, and TL were supported by
MARS (Managing Aquatic ecosystems and water Resources
under multiple Stress) funded under the 7th EU Framework
Programme (Contract No.: 603378). JR was supported by
the German Ministry of Education and Research (project
NITROLIMIT, grant no. 033L041 A).
ACKNOWLEDGMENTS
We thank all technicians of IGB Berlin responsible for the
long-term measuring program in Lake Müggelsee. Antje Barsch,
Nadine Baadke (Landesumweltamt Brandenburg) and Antje
Köhler (Senat Berlin) provided data for selected German lakes.
The Environment Agency provided water quality data and the
Broads Authority macrophyte data for the lakes in the Broads
National Park (UK). We acknowledge linguistic improvements
by Anne Mette Poulsen. We thank two reviewers for their helpful
comments.
SUPPLEMENTARY MATERIAL
The Supplementary Material for this article can be found
online at: https://www.frontiersin.org/articles/10.3389/fpls.2018.
00194/full#supplementary-material
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Conflict of Interest Statement: The authors declare that the research was
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The handling Editor is currently co-organizing a Research Topic with one of
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Copyright © 2018 Hilt, Alirangues Nuñez, Bakker, Blindow, Davidson, Gillefalk,
Hansson, Janse, Janssen, Jeppesen, Kabus, Kelly, Köhler, Lauridsen, Mooij,
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Donk, Waterstraat, Willby and Sayer. This is an open-access article distributed
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