PhD thesis - ku Schou Frandsen.pdf · PhD thesis Mette Cristine Schou Frandsen ... Over the course of three years I followed the small scale variation in the natural tracer,

PhD summary F A C U L T Y O F S C I E N C E

U N I V E R S I T Y O F C O P E N H A G E N

PhD thesis

Mette Cristine Schou Frandsen

Ecohydrological investigations of a groundwater- lake system

- A cross disciplinary study in the interactions between biology, lake ecology and hydrology.

Academic advisor: Ole Pedersen

Peter Engesgaard

Bertel Nillson

Submitted: 31/10/14

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Data Sheet

This PhD thesis has been submitted to the PhD school of the faculty of Science at University of

Copenhagen.

Name of department: Department of Biology

Author: Mette Cristine Schou Frandsen

Title / Subtitle: Ecohydrological investigations of a groundwater- lake system

- A cross disciplinary study in the interactions between biology, lake ecology

and hydrology.

Academic advisors: Ole Pedersen

Department of Biology

University of Copenhagen

Peter Engesgaard

Department of Geosciences and Natural resource Management

University of Copenhagen

Bertel Nilsson

Department of Hydrology

Geological Survey of Denmark and Greenland

Submission: Oct 31 2014

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Preface

This thesis concludes a three years PhD program in collaboration between the Department of Biology (DB),

University of Copenhagen; the department of Geosciences and Natural Resource management (DGN),

University of Copenhagen; and Denmark and Greenland Geological Survey (GEUS).

The PhD project was founded by The Danish Council for Independent Research – Nature and Universe.

The study was supervised by Associate Professor Ole Pedersen (DB), Professor Peter Engesgaard (DGN)

and senior scientist Bertel Nilsson (GEUS). An external research stay of three month was spent at the

School of Earth and Environment, University of Western Australia, Crawly, WA, Australia under the

supervision of Professor Matthew R. Hipsey.

In accordance with the guidelines given by the faculty of Science, University of Copenhagen, this thesis

consists of a summary and the following 4 papers:

Paper I: Frandsen, M,. Nilsson, B., Engesgaard, P., Pedersen, O. 2012. Groundwater Seepage

stimulates the growth of aquatic macrophytes. Freshwater Biology. 57:907-921.

Paper II: Frandsen, M,. Engesgaard, P., Nilsson, B., Pedersen, O. Rooted underwater vegetation

locally reduces groundwater discharge in lakes.

Paper III

Frandsen, M,. Engesgaard, P., Nilsson, B., Pedersen, O. Tacking groundwater flow during a flow

reversal – nature’s own tracer experiment.

Paper IIII:

Frandsen, M,. Engesgaard, P., Nilsson, B., Pedersen, O. Using whole-system understanding to

evaluate long term development in alkalinity in a northern flow through lake.

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Acknowledgements

First and foremost, a special thanks to my supervisors Peter, Bertel and Ole. To Peter Engessgaard, for always being there to help, no matter what the subject might be, for always giving me constructive feedback, for always offering his knowledge, for all the many, and often cold hours in the field, and all the good evenings at the field station playing ping pong. A special thanks to Bertel Nilsson for his kindness and his constructive feedback and for his valuable help on setting up the monitoring equipment during some very cold December days. A special thanks to Ole Pedersen, for his always constructive and detailed feedback, for taking care of many of the practicalities, and for making sure that I stayed on the right course. A big thanks to Heidi Barlebo. for her kind support, and for always seeing the positive side. A big thanks to Matthew R. Hipsey for his kind support during my stationing at University of Western Australia in Perth. I would also like to give a big thank to Carlos Duque Calvache and Mikkel Rene Andersen for their valuable help in the field. A big thank to Mitra Christin Hajati and Kristian Färkkilä Knudsen for contributing with data. I would also like to thank colleges and students at the Geological Survey of Denmark and Greenland, the Freshwater biological laboratory, department of Geosciences and Natural resource Management at the University of Copenhagen, and The University of Southern Denmark for their contributions. Per Jørgensen and Jens Bisgaard are thanked for their technical assistance. A warm and deep thank to my twin sister Marie Michelle Schou Frandsen, for always being there for me, no matter the cost. For helping me through the rough times with kindness and love, and for always believing in me. The biggest thanks go to Thomas Duus Henriksen, for being there for me, for believing in me, for keeping my spirit up, for supplying sweets and food in the late hours, for going all the way to Australia with me and for making the study possible. This project was founded by The Danish Council for Independent Research – Nature and Universe, who I would like to thank for granting me the opportunity to conduct this PhD study.

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Preface

Acknowledgements

1 Introduction and objectives ....................................................................................................................... 5

1.1 Background .............................................................................................................................................. 5

1.2 Motivation and objectives ....................................................................................................................... 7

2 PhD Research ............................................................................................................................................ 8

2.1 Paper I ...................................................................................................................................................... 9

2.1.1 Introduction and objectives ................................................................................................................. 9

2.1.2 Main findings ........................................................................................................................................ 9

2.2 Paper 2 ................................................................................................................................................... 10

2.2.1 Introduction and objectives ............................................................................................................... 10

2.2.2 Main findings ...................................................................................................................................... 11

2.3. Paper 3 .................................................................................................................................................. 11


2.3.2 Main findings ...................................................................................................................................... 12

2.4. Paper 4 .................................................................................................................................................. 12


2.4.2 Main findings ...................................................................................................................................... 13

3 Conclusions and perspectives .................................................................................................................. 13

4 References .............................................................................................................................................. 15

Appendixes .................................................................................................................................................. 1

Paper 1 Groundwater Seepage stimulates the growth of aquatic macrophytes

Paper 2 Rooted underwater vegetation locally reduces groundwater discharge in lakes

Paper 3 Tacking groundwater flow during a flow reversal – nature’s own tracer experiment

Paper 4 Using whole-system understanding to evaluate long term development in

alkalinity in a northern flow through lake

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Abstract

This PhD project is a cross-disciplinary study combining hydrological and biological methodology to better

describes the lake-catchment interaction seen in an ecological perspective.

The topics investigated were:

Does groundwater re- and discharge affect the growth of submerged vegetation? (Paper I).

Does dense bottom vegetation affect the small scale hydrology of the lake bed sediment? (Paper 2).

How can natural tracers (δ 18

O) be used to quantify the temporal variation in groundwater seepage dynamics? (Paper 3).

Is it possible to combine ecological data of surface water chemistry and data on groundwater chemistry to stoichiometrically describe changes in the lake in a historical time frame? (Paper 4).

The main conclusions from the study are:

When evaluating the ecology of a groundwater-lake system, both hydrological and biological parameters are needed to accurately describe the factors affecting the system.

The biology and ecology of the lake (i.e. submerged vegetation and surface water chemistry) are highly affected by groundwater seepage.

The hydrology at the surface-water-interface is highly affected by the biology (i.e. submerged vegetation).

Groundwater-lake systems are very dynamic systems on a spatial scale. Variability in meteorology can lead to variability in the hydrology, and in some cases ignite transient effects that are temporally distinct and difficult to capture.

To some extend the lakes acts as sentinel for all the in and out-puts to the system as well as the in-lake processes. By combining this ecological view with hydrology, it is possible to gain information on the historical development in the surface water chemistry.

Lake Hampen is a Danish flow through lake receiving almost 2/3 of its water through groundwater discharge.

In this setting I investigated the interrelationship between hydrology and biology.

I found that groundwater seepage significantly affected the growth rates of submerged isoetids

(small rosette type plants) by providing them with a continuous supply of nutrients and inorganic carbon. The

seepage rates were strongly correlated to the growth responses and the plant mass was higher in treatments

where the plants were subjected to groundwater seepage compared to treatments with no groundwater

seepage.

I also found that the submerged vegetation conversely had a significant effect on the small scale hydrology

of the lake bed sediment. On densely vegetated areas (~9000 plants m-2

), the vertical hydraulic conductivity

was lower compared to non-vegetated sediment. Disturbing the top layer of the sediment lead to a significant

increase in hydraulic conductivity on the vegetated sediment, whereas the non-vegetated sediment was not

affected by this. The reasons for the lowered hydraulic conductivity seems to be an combination of the

organic content in the sediment (i.e. the roots of the plants) and a vegetation induced entrapment of fine

particles in the sediment.

Over the course of three years I followed the small scale variation in the natural tracer, δ18

O, and

nitrate in the main discharging area of the lake to follow an ongoing flow reversal in the system. By tracking

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the groundwater – lake water signal using only the distribution of δ18

O it was clear that lake water had

penetrated the lake bed sediment down to at least 1.25 m during the flow reversal. This was also clear

looking at the nitrate data and during the flow reversal the nitrate concentrations in the sediment was

significantly lower than under normal flow conditions. All the nitrate was denitrified before reaching the lake

and the estimated denitrification rates were lower than the assumed capacity.

In Lake Hampen, the alkalinity suddenly started to increase during the mid-1970s. Using a simple

four step modeling approach, I found that denitrification of nitrate discharging to the lake, stoichiometrically

could explain the development in the alkalinity in the surface water. This method gave a surprisingly accurate

picture of the yearly development in surface water alkalinity despite the somewhat simplified approach used

to estimate the historical input of nitrate with the groundwater.

In conclusion I strongly encompass the notion that a cross-disciplinary approach greatly qualifies the

results of ecological studies.

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Dansk résumé

Dette PhD projekt er et tværfagligt studie hvori metodik fra hydrologi og biologi er blevet brugt til at lave

tværfaglige økologiske undersøgelser. I projektet undersøges følgende emner:

Påvirker grundvandsindsivning i søer vandplanterne? (artikel 1).

Påvirker vandplanterne de hydrologiske forhold i søbunden? (artikel 2)

Kan man bruge naturlige tracers til at kvantificere den tidslige variation grundvandsdynamikken?

(artikel 3)

Kan man ved at kombinere økologiske data og hydrologiske data beskrive den historiske udvikling i

en sø?

Hovedkonklusionerne fra projektet er:

Når man skal evaluere økologien af et grundvand-søsystem, er der brug for både hydrologiske og

biologiske data hvis man vil vurdere de faktorer der påvirker systemet.

Vandplanterne er signifikant påvirket af grundvandsindsivningen.

Vandplanterne har en stor effekt på de hydrologiske forhold I søbunden, og påvirker derfor

grundvandsstrømningen til systemet.

Grundvand-søsystemer er meget dynamiske systemer bade spatialt og temporalt. Variabilitet i

meteorologi og hydrologi, kan i nogle tilfælde skabe midlertidige ændringer i grundvandet

strømningsmønster.

Overflade vandet i en sø indeholder informationer om alle til- og fraførsler af stoffer samt de interne

processer. Ved at kombinere hydrologiske og biologiske data kan man bruge overfladevandets

”hukommelse ” til at forklare den historiske udvikling i søkemien.

Hampen sø er en dansk grundvandspåvirket sø, som modtager næsten 2/3 af sit vand fra grundvandet. På

denne lokalitet undersøgte jeg koblingen mellem hydrologi og biologi.

Gennem studiet fandt jeg at grundvandsindsivning havde en signifikant indflydelse på vækstrater for

undervandsvegetation ved at forsyne denne med en vedvarende forsyning af næringsstoffer.

Vækstraterne var stærkt korrelerede til grundvandsraterne og planterne der blev udsat for grundvand

i forsøgene opnåede en større slutmasse end planter fra forsøg hvor grundvandtilførslen blev afskåret.

Undervandsvegetationen havde stor effekt på de hydrologiske forhold i søbunden. På tæt

bevoksede søsedimenter kan der vokse op til 9000 planter per kvadratmeter. Denne Undervandsvegetation

havde stor effekt på de hydrologiske forhold i søbunden. Ved at lave sammenlignende studier af henholdsvis

bevokset og ubevokset sediment, kunne man se at vegetationen skabte et lag af lav hydraulisk

ledningsevne. Årsagen til den sænkede hydrauliske ledningsevne syntes at stamme fra en kombination af

det organiske indhold i sedimentet (eksempelvis planterødderne), og en vegetationsskabt indfangning af

småpartikler i sedimentet.

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I løbet af en treårig periode fulgte jeg δ18

O fordelingen i sedimentet i et forsøg på at undersøge en

retningsændring af grundvandet (et flowreversal). I et område hvor der normalt indstrømmer store mængder

nitrat med grundvandet, løb grundvandet i stedet ud af søen. Ved at undersøge fordelingen af δ18

O blev det

klart søvandet i løbet af dette flow-reversal havde trængt mindst 1,25 meter ned gennem

søbundssedimentet. Dette fremgik tilsvarende tydeligt af nitratkoncentrationerne i sedimentet. Under dette

flow-reversal var nitratkoncentrationerne i sedimentet væsentligt lavere end de var under normale

omstændigheder.

I Hampen Sø begyndte alkaliniteten at stige op gennem midten af 1970erne. Ved at anvende en

simpel, 4-trins modelleringstilgang fandt jeg at denitrifikationen af nitrat der tilstrømmer søen med

grundvandet støkiometrisk kunne forklare alkalinitetsudviklingen i overfladevandet. På trods af sin simple

tilgang gav denne metode et overraskende præcist billede af den historiske alkalinitetsudvikling i

overfladevandet og alakalinitetstigningerne i sen kan kobles til udledningen af nitrat fra et landbrug der tæt

ved søen.

I sin konklusion tilslutter projektet sig på det kraftigste ideen om at en tværdisciplinær tilgang ikke

kun kvalificerer økologiske studier, men også er en nødvendighed for at kunne tegne et realistiske billede af

de processer der styrer og påvirker søøkologien i grundvandspåvirkede søer.

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1 Introduction and objectives

1.1 Background

Freshwater lakes are extremely important to us as they provide us with water for domestic, agricultural and

industrial use and in some regions act as drinking water reservoirs for both humans and animals (Brønmark

& Hansson 2002). They also play a key role in maintenance of the ecosystems and the species diversity of

plants and animals (Brønmark & Hansson 2002; Zalewski, 2000). The qualitie of the lakes are threatened by

numerous factors, the more important ones on the northern hemisphere being pollution and eutrophication

(Brønmark & Hansson 2002: Brinson & Malvarez 2002). Eutrophication caused by nitrate leaching from

agriculture and phosphorus leaching from populated areas represent one of the major anthropogenic threats

to freshwater lakes in Europe (EEA, 2005). In countries as Denmark and the Netherlands, nitrate leaching

from agriculture poses a major problem as more than 60% of the areal surface in these countries are

cultivated for agriculture (Schaap et al., 2011; Hansen et al., 2012). Hence, one of the main goals in water

management is to handle and find solutions for environmental problems in order to obtain and maintain

favorable ecological status in wetlands (i.e. lakes, streams).

The problems with Lake eutrophication became evident during the 1950s and 1960s where many

lakes in agricultural and urban areas experienced algal blooms, fish death kills and deterioration of

submerged vegetation (Brønmark & Hansson, 2002). In the 1960s, only few regulations regarding storage

and disposal of industrial wastes, fuels, chemicals and fertilization of cropped fields existed. Unregulated

amounts of harmful substances were released to the groundwater all over the industrialized part of the world

(Brønmark & Hansson, 2002). In 1974, Schindler conducted one of the first whole lake experiments directly

linking nutrient concentrations in the lakes to the observed problems (Schindler, 1974). Subsequently

numerous regulations have been posed trying to stop the pollution and eutrophication of our freshwater

systems, i.e. the Danish NPO-regulation and the European habitat directory (HD), the Natura-2000 plans and

finally the Water Framework Directory (WFD), which states that good ecological status must be achieved by

2015 in all water bodies (European Union 2000). These regulations have had a positive effect on especially

the nitrate leaching from agriculture which has been reduced by 50% between 1990 and 2003 (Blicher-

Mathiasen et al. 2013; Wiberg-Larsen 2013). However, despite these attempts to address the problem, many

of the freshwater environments are still deteriorating (Danish Nature Agency, 2014; Kundzewicz, 1999).

First of all, there are problems restoring already damaged wetlands. In the mid-1980s and the 1990s,

it was established that simply reducing the load of nutrients to a lake, was insufficient to restore it to good

conditions as biotic feedback mechanism captures the lake in poor condition (Timms & Moss 1984; Brock

and Starrett, 2003). In short, a lake can exist in two states under the same nutrient concentration. One is

dominated by phytoplankton resulting in turbid water, and one dominated by submerged plants, with clear

water (Timms & Moss 1984; Scheffer 1990). Even though changes in nutrient load can shift a lake from the

clear water state to the turbid state, the reversed process are more difficult due to the biotic buffer systems.

Restoring a lake in bad condition is very costly and the restoration methods are invasive and difficult to test.

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However, this is a problem being addressed in the movement in Demark by a Centre of Excellency funded

by the Villum Foundation. In this program scientists from multiple disciplinary areas tests and evaluates the

different restoration methods (CLEAR, 2006).

Another great challenge in maintaining good conditions in our surface waters, is of more academic

character. Lake management and lake ecology are traditionally biological disciplines, while groundwater flow

and transport in the catchment sediment are traditionally geological and hydrological disciplines. Hence,

coupling hydrology and ecology crosses well established disciplinary borders, often resulting in studies

lacking either the hydrological inputs or the ecological response (Zalewski, 2002).

On a catchment scale the groundwater surface water systems are complexes of interrelated hydrological,

geological, biological, and chemical processes, and in order to understand the impact on a wetland from the

surrounding catchment the hydrological framework is needed and vice versa. While hydrologist seems to be

embracing the new discipline of ecohydrology, it has been suggested that biologist are less aware of this

new emerging discipline (Bond, 2003). Hannah et al. (2004) state in their bibliographical analysis of the

emergence of ecohydrology, that of the articles published from 1981-2004 using the word ecohydrology

(hydroecology, hydro ecology or eco hydrology) 71% appeared in physical journals whereas only 23%

appeared in biological journals. This does, however, only accounts for papers using those specific words.

Thus, their study likely underestimated the overall extend of the ecohydrological research, as they also state

themselves. Furthermore, even though it seems that hydrologists are more actively involved in the new field

of ecohydrology, a literature review showed that despite the use of the word ecohydrology in the hydrological

publications and the apparent focus on ecological implications in the systems they describe, they often lack

the biological data to support it (Hannh et al.2004).

For both hydrologist and biologist/ecologist, the search for solutions to practical problems has played

a central role in the development of these fields, and the advancements have been driven by innovation in

research techniques making it possible to address these problems (Kundzewicz, 2000; Groffman & Pace,

1998; Nuttle 2002). While this has led to a good mechanical understanding on how both the hydrological and

the ecological systems work, we still need to combine this knowledge into a more holistic understanding of

the ecosystems (Zalewski 2002). Zalewski states that the field of ecohydrology is the third fase in the

development of ecology. The development starts with a pure descriptive natural history (ie. Linné), followed

by an understanding of the processes within the system, which finally leads to an understanding that makes

it possible to control and manipulate the system to ensure and secure resource quality and availability

(Zalewski, 2002). The overall goal from my point of view is to develop a scientific framework enabling us to

develop the full set of skills necessary to implement sustainable management of natural resources. Here,

aquatic eco-sciences are important. In 1998, the World Science Report (UNESCO, 1998) stated that

protection of our water resources, in the face of increasing deterioration of the global environment, is one of

the priority goals for science. Water resources can be sustained, not only by reducing treads, but also by

regulation the system within the drainage basin. To do so, we need to integrate and combine methodology

and knowledge from both hydrology and ecology in ecohydrology (Zalewski et al., 1997, Zalewski, 2000).

In lake ecology focus has been on surface water chemistry and physical conditions within the lakes,

whereas the groundwater re- and discharging in these lakes to a large extend represent an understudied

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field in lake ecology. Several studies show that groundwater seepage can contribute with up to 50% of the

annual nutrient load to a lake (Brock et al. 1982; Shaw and Prepas, 1989; Sebestyen and Schneider, 2004;

Cullmann et al., 2005). Ommen et al. (2012) showed the same for a Danish seepage lake where they

estimated groundwater seepage accounted for 67% of the total P input and 50% of the total N input. This will

undoubtedly affect the lake ecology and with this project, I aim to address some of the interactions between

hydrology and ecology in a seepage lake.

1.2 Motivation and objectives

During the past 100 years agricultural production in Denmark has increased significantly (Hansen et al.,

2012). Following World War II, the international trade with fertilizers and feed boomed, and in Denmark the

import of both virtually exploded during the sixties to eighties (Hansen et al,. 2012). This was reflected in the

nitrate concentrations in the Danish groundwater and a clear increasing trend was observed following the

increased use of fertilizers (Hansen et al., 2012). Despite the significant reduction in the use of fertilizers

following the regulations described above, the environmental goals to meet the requirements of the Water

Framework Directive for the majority of the Danish lakes are not yet reached (Danish Nature Agency, 2014).

In order to effectively address the above problem, there is a need for better understanding of how

the individual water systems are impacted. In Denmark 80 restoration attempts have been done within the

past 20 years. In many cases the effect of the restoration decrease after just a few years. This is often

attributed to either internal loading of phosphorous in the system or poor control of nutrient input from drains

and surface inlets (Bramm and Christensen, 2006). Surprisingly, it seems that in many of the cases, no

attempts to quantify nutrient input through diffuse sources such as groundwater discharge have been

considered (Søndergaard et al,. 1999; Liboriussen et al., 2007). This is, however a very plausible explanation

for why some of the lakes have returned to the bad condition it was in before the restoration initiative.

In this PhD study a groundwater/lake system that is largely impacted by terrestrial inputs from both

agriculture and forested land (Kidmose et al., 2011; Ommen et al., 2012, Karan et al 2014) is investigated

from an ecohydrological perspective. Using small scale experimentally derived data it is investigated how

groundwater affects the biology, and conversely how the biology affects the hydrological properties of the

sediments and the groundwater chemistry at the surface water interface (SWI). On a larger scale it is

attempted to incorporate these small scale hydrological, chemical and biological processes in a whole-lake-

response model.

The main objectives were to 1) evaluate how the groundwater seepage affected the growth of

different types of submerged plants. 2) Investigate how the submerged vegetation affects the small scale

hydrology in the lake bed sediments. 3) Investigate the small scale flow patterns and transport of nitrate

through the groundwater by using natural tracers. 4) Use data on different temporal and spatial scales, to

evaluate a historical ecological effect of the nitrate leaching to the lake.

Addressing these objectives has resulted in four papers:

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Paper I, “Groundwater Seepage stimulates the growth of aquatic macrophytes” addresses objective 1 and is

published in Freshwater Biology.

Paper II, “Rooted underwater vegetation locally reduces groundwater discharge in lakes” addresses

objective 2.

Paper III “Tacking groundwater flow during a flow reversal – nature’s own tracer experiment” addresses

objective 3.

Paper IIII “Using whole-system understanding to evaluate long term development in alkalinity in a northern

flow through lake” addresses objective 4.

2 PhD Research

All experiments were conducted in Lake Hampen, Denmark. Lake Hampen is a well investigated lake

(Moeslund, 2000; Kidmose et al. 2011, Ommen et al., 2012; Karan et al. 2014) which allows the results from

this study to be compared with previous findings and gives access to historical data on the lake surface

water chemistry. Lake Hampen is characterized as a flow-through-lake and approximately 2/3 of the water is

received from ground water seepage discharge at the north-eastern side of the lake and approximately 3/4

leaves the lake through groundwater seepage recharge at its westerns side (Ommen et al., 2012). These

settings are ideal for studying the interactions between hydrology and lake ecology.

Paper 1 addresses the effects of groundwater seepage discharge on two species of submerged

plants with contrasting morphological adaptation for nutrient and carbon uptake. Paper 2 addresses how the

carpet like structures of isoetids can affect the small scale hydrology by forming a layer of low hydraulic

conductivity. Paper 3 uses a natural tracer (δ18O) to follow movement of groundwater and nitrate to and

from the lake during a flow reversal where the directional flow of the groundwater changed. Under normal

conditions, groundwater discharges large amounts of nitrate to the lake, but during the flow reversal, lake

water penetrated the sediment down to at least 1.25 m. Paper 4 addresses the historical effects of

groundwater seepage discharge on the lake ecology by investigating how the groundwater input of nitrate

may have affected the historical development in alkalinity of the lake water

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2.1 Paper I

Title: “Groundwater seepage stimulates the growth of aquatic macrophytes”

2.1.1 Introduction and objectives

Submerged aquatic plants are a diverse group of organism competing for the same resources (Carignan &

Kalff, 1980; Rattray et al., 2991). Whereas most plants can utilize the inorganic carbon and nutrient from the

water column, only specially adapted plants are able to use the sediment as a source of these substances

(Sand-Jensen & Prahl, 1982; Madsen et al., 2002). In northern temperate lakes, the availability of inorganic

carbon and nutrient have considerable seasonal variation with low concentration in the growth season where

nutrients are bound in living biomass, and much higher concentrations during winter when mineralization

processes dominate (Guilford & Hecky, 2000). Plants that are able to exploit the nutrients in interstitial water

in the sediment have a competitive advantage during the growth season (Madsen et al. 2002).

The availability of inorganic carbon and nutrients are normally evaluated by measuring the

concentrations in the surface water and the interstitial water. The groundwater represents a largely

overlooked source of both nutrients and inorganic carbon. Groundwater contains high concentrations of

inorganic carbon due to subsurface respiration processes and it is often relatively nutrient rich due to

accumulation in the catchment (Brock et al. 1982; Shaw & Prepas, 1989; Hagerthey and Kerfoot, 1998). In

seepage discharge zones, there is a continuous supply, and this could have large effects on the vegetation

benefiting from this source. How this groundwater affects the submerged vegetation is poorly understood.

Some studies suggest that it might have an effect, but to my knowledge no studies examines this directly.

Actual growth experiments have been carried out on algae by Hagarthy and Kerfoot (1998) and they find a

significant effect of groundwater seepage on the growth rates of the algae.

The objectives of this study was to 1) Examine the effects of re- and discharge from catchments with

different land use on the growth rates of submerged plants. 2) Examine the in situ and in vitro effect of

groundwater seepage on plants of contrasting morphological adaptions to nutrient and carbon uptake. 3)

Examine the relative importance of groundwater seepage discharge as a nutrient source, a source of

inorganic carbon, or both.

2.1.2 Main findings

We found that groundwater discharge had a significant, positive effect on the in situ growth rates of Littorella

uniflora. This was evident on both an area where the groundwater originated from a forested catchment and

from an area where the groundwater originated an agricultural catchment. The plants benefited both from the

inorganic carbon and the nutrients (nitrate of phosphorous), but the results indicated that mainly the

increased availability of inorganic carbon resulted in the enhanced growth. The final plant mass was up to

70% higher in plants that were subjected to seepage discharge when compared to the control plants who

received no groundwater.

We found a strong positive correlation between seepage discharge rates and the growth rates and

final plant mass. This was evident on both the areas where the discharge water originated from a forested

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catchment and the area where the discharge water originated from an agricultural catchment, but not on

plants grown in an area of groundwater recharge.

2.2 Paper 2

Title: “Rooted underwater vegetation locally reduces groundwater discharge in lakes”


It is generally accepted that groundwater discharge in homogeneous systems mainly takes place near the

shore and decreases exponentially with distance to shore (McBride & Phannkuck, 1975; Lee, 1977; Shaw &

Prepas. 1990). The groundwater discharging to the lake percolates the littoral zone in the near shore areas.

Here the water must surpass the roots of the submerged vegetation, before entering the lake surface water.

Some species of submerged plants called isoetids form carpet like structure holding ~9.000 individual plants

m-2

(Christensen and Sørensen, 1986). These plants have special morphological adaptations that give them

an effective nutrient uptake through their roots that often comprise up to 50% of the total plant mass

(Hutchinson, 1975). Due to their large root systems, the isoetids may effectively alter the hydraulic properties

of the near shore sediment.

The roots can have a direct effect on the hydraulic conductivity, simply by filling up pore spaces or by

altering the overall porosity of the sediment. They can also indirectly affect the hydraulic conductivity by

promoting intrusion of fine particles in the sediment (Lehman 1975, Hilton 1985). The canopy of submerged

vegetation attenuates the energy associated with waves and currents herby creating a low energy

environment that allows entrapment of fine particles (Hilton, 1985; Blais & Kalff, 1995). The intrusion of fine

particles (≤ 63µm) reduces the sediment porosity, herby lowering the hydraulic conductivity (Lehman, 1975;

Hilton, 1985). This phenomenon is well studied in stream settings (Brunke & Gonser, 1997; Huettel et. al.,

1996; Huettel & Gust, 1992), but has not been addressed for lake settings.

However some studies suggest the existence of a vegetation induced lowering of the hydraulic

conductivity. Frandsen et al. (2012) and Hargerthey & Kerfoot (1998) both found an interesting increase the

seepage discharge after installing similar growth chambers in lake bed sediments. In both studies, it is

speculated whether a layer of low hydraulic conductivity is punctured during the installation of the growth

chambers. A layer of low hydraulic conductivity caused by vegetation was also suggested of Karan et al.

(2014). They found evidence of lowered hydraulic conductivity in the littoral zone, and identified an off-shore

discharge peak of groundwater, which could only be explained by including a lower permeable lake bed in

the near shore area in their model. They speculate if the vegetation might divert the groundwater flow, herby

explaining the off-shore groundwater discharge peak.

The objectives of this study were to examine the relationship between hydraulic conductivity and

small-scale sedimentary conditions caused by the presence of vegetation by comparing results from both a

vegetated and a non-vegetated area of a lake bed in a groundwater-fed lake.

PhD summary

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2.2.2 Main findings

We found evidence for a strong vegetation induced effect on the hydraulic conductivity in the lake sediment.

Based on my results, this was caused by a combined effect of organic content in the sediment and the mass

of fine particles (<63µm).

We conducted simple disturbance experiment on both vegetated and non-vegetated sediment. On the

vegetated sediment, the disturbance had a strong effect and the hydraulic conductivity increased

significantly. In comparison, the hydraulic conductivity was not affected significantly on the bare sediment.

The relative mass of fine particles (<63µm) was significantly higher on the vegetated area, compared

to the non-vegetated area. The lowest hydraulic conductivity was found in the same depth normally identified

with the highest root density. We found a strong correlation between organic matter and mass of fine

particles in the sediment, and we speculate that the roots might trap and hold the fine particles in the

sediment.

2.3. Paper 3

Title: “Tracking groundwater flow during a flow reversal – nature’s own tracer experiment”


The interaction between groundwater and surface water is complex and difficult to accurately quantify. The

spatial variation makes it costly and time consuming to map the in- and outputs correctly, and it is often not

possible. On a large scale, the groundwater systems are controlled by recharge from precipitation, drainage

through discharge to surface water, evaporation, and evapotranspiration. The groundwater flow pattern is

mainly controlled by topography (gravity) and flows from high to low elevations (Winter, 1999). On a smaller

scale, the flow pattern is affected by sedimentary heterogeneity, which causes differences in hydraulic

conductivity. Furthermore, the temporal variation in the groundwater flow pattern can vary from hours to

decades, which makes the system very dynamic in both time and space.

Locally, the flow system associated with surface water bodies results in more complex flow patterns,

regardless of topography. Seasonal changes in lake stage, transpiration, evaporation and precipitation might

change the head differences between the lake and the groundwater and have substantial effect on the

groundwater – lake interactions. The sensitivity of the groundwater-surface water flow systems gives rise to a

number of transient effects (Cheng and Anderson, 1994, Anderson and Cheng, 1993). Of the more important

transient effects is the flow reversal. Flow reversal alters the directional flow of ground water, sometimes

changing recharge areas to discharge areas. In ground water - lake systems where the groundwater

transport large amount of nutrients to the lakes, this may affect the ecology of the lake (Sacks et al. 1992)

The objectives of this study were to quantify the duration of an ongoing flow reversal at Lake Hampen by

using data on the natural δ 18

O isotope, and at the same time examine the effect of the flow reversal on the

nitrate transport to the lake during the reversal.

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2.3.2 Main findings

During this study the back seeping groundwater after a flow reversal in Lake Hampen was captured. During

the flow reversal, lake water penetrated the lake sediment down to a depth of more than 1.25 m and during

2011 – 2013 the groundwater gradually returned. The results suggest that the flow reversal has been in

effect from at least 2010 to ~2012-2013.

The flow reversal was probably caused by meteorological factors. The lake stage and groundwater head

were low in 2010 and 2011. The total precipitation decreased from 913 mm in 2007 and 875 mm in 2008 to a

mere 754 and 704 in 2009 and 2010 respectively. This could have affected both the lake stage directly, but

also the groundwater head, due to lower infiltration (Downing and Peterka, 1978; Sacks et al., 1992;

Rosenberry et al., 1997). At the same time, the air temperature during the winter dropped in 2009 and 2010,

causing a larger proportion of the precipitation in those years to fall as snow, possibly causing a lower

infiltration rate.

It is speculated that none of the meteorological data is very extreme seen in a historical time frame, and if

these factors leads to a groundwater reversal in Lake Hampen, it should be assumed that the lake most

likely has undergone several groundwater reversals historically.

2.4. Paper 4

Title: “Using whole-system understanding to evaluate long term development in alkalinity in a northern flow

through lake”


The alkalinity in Lake Hampen has increased from the mid-1970s until the mid-1990s. The reason for this is

unknown, but some studies suggest that denitrification of nitrate seeping to the lake, from an agriculture area

on the North-East shoreline, might be a possible source of alkalinity (Karan et al. 2014). The in-lake alkalinity

production is driven by, on the one hand primary production, and on the other hand reduction of the major

anions such as sulfate and nitrate (Schindler 1986; Rudd et al. 1988). Studies show that these processes

can play a significant role in the regulation of the acid-base system, especially in soft water lakes (Davidson,

1986). In the past ~80 years, the nutrient load to lakes has increased and as a consequence a general trend

of increased alkalinity has been observed in many of the lakes (Sutcliffe et al., 1982).

Nitrogen discharge in the form of nitrate can be especially harmful to the lakes. The anaerobic denitrifiers

need nitrogen in the form of NO3- and this is often supplied from the aerobic nitrifying processes where

NH4+ (ammonium) is oxidized to NO2-/NO3- (Appelo & Postma, 2005). However, if the nitrogen is readily

supplied in the form of nitrate, the denitrifiers only need a sufficient electron donor (i.e. carbon, pyrite or

´Fe2+) to drive denitrification.

The objective of this study was to examine if agriculturally derived nitrate discharging to the lake form the

agriculture could explain the changes in alkalinity from the mid-1970s till today. In the absence of historical

data on the nitrate concentration in the groundwater, a four step mass balance approach is used to

investigate this. Information on the surface water concentrations of nitrate, EC and alkalinity and a correlation

PhD summary

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between nitrate and EC in the groundwater was used to estimate the historical input of nitrate. I then tried to

denitrify the excess nitrate given by the mass balance model to see if this would give rise to alkalinity

changes as the ones we observe in the lake.

2.4.2 Main findings

Using long time series of data on surface water chemistry dating back to the 1970s and a simple mass

balance approach I show that denitrification of nitrate-polluted groundwater coming from the agricultural part

of the catchment could be an explanation for the observed increase in alkalinity in the lake during the 1980s

and 1990s.

Using a simple mass balance approach, it was possible to predict the alkalinity changes in Lake Hampen in a

historical time frame surprisingly accurate. Given the assumptions used, denitrification of nitrate discharging

into the lake from the agriculture could stoichiometrically account for the observed changes in alkalinity. The

denitrification rates used are in good agreement with what other studies have found in the lake.

3 Conclusions and perspectives

The discipline of ecohydrology is called a new and emerging discipline (Hannah et al., 2004). It is difficult to

conduct true cross-disciplinary studies, as scientists are normally only trained thoroughly in their own main

discipline. Still, given the number of challenges standing before us in regard to maintaining our freshwater

systems, the need for cross-disciplinarity has never been more urgent. The underlying objective of this study

was to examine if cross-disciplinary methods would strengthen the way we describe and investigate our

freshwater lake systems as either biologist or hydrologist. I have tried to include the hydrological

methodology in the traditional ways we do biological studies and vice versa.

In the first paper, I show how the groundwater discharging into the lake through the lake bed sediment have

a significant effect on the growth rates of the submerged plants (Frandsen et al., 2012). This is a significant

finding. Looking at the literature, it seems that seepage discharge of nutrients to the submerged plants is

almost completely overlooked. In example Barko et al. (1991) review the interactions between sediment and

submerged plants. They evaluate the sources of particular nutrients for uptake by submerged plants without

mentioning groundwater seepage discharge as a possible source. Few studies include this source of

nutrients (Frandsen et al., 2012; Carpenter and Lodge, 1986; Sebestyen and Schneider, 2004). In lakes

where groundwater discharges, the plant communities have a continuous supply of nutrients and carbon

from the groundwater, and this might be a significant factor driving growth and possibly also distribution. This

should be included in future in situ studies of submerged plant growth and distribution.

Furthermore, as the groundwater seepage has a positive effect on the growth rates of the

submerged isoetids, it also means that the isoetids filters the water from nutrient and carbon before it enters

the surface water. This is also interesting as the thick carpet like structures of isoetids might play a key role

in preventing eutrophication in some systems. Ommen et al. (2012) have estimated that the isoetids in Lake

Hampen might take up as much as 1.7 ton N year-1

that would otherwise have ended up in the surface water.

PhD summary

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Paper two addresses how the isoetids vegetation affects the small scale hydrology of the lake bed

sediment by lowering the hydraulic conductivity. This is also an interesting finding. First of all, it could lead to

an increased discharge further away from the shore to off-shore areas with no effective plant cover to filter

the water (Karan et al. 2014). From a lake management perspective, diverted groundwater seepage could

represent an undetected source of nutrients as these off-shore discharge zones can be difficult to detect.

Hence, these findings from paper 1 and paper 2 give new examples on the often high degree of uncertainty

up-scaling point observations to whole system analysis often are succumbed to. If the nutrient content in the

pore water is used to conclude the overall nutrient availability in the lake, it will be greatly underestimated in

lakes where the groundwater seepage transports large amount of nutrient to the system. Even when this is

taken into account, off shore peaks in groundwater seepage, could be overlooked.

The problem of up-scaling point observations to whole system analysis can, however, be overcome.

To estimate the overall effect of groundwater on a lake, knowledge is needed on both the exchange of water

and the exchange of solutes through groundwater re- and discharge. However, it is not possible to do a

complete mapping of either as both the spatial and temporal variations are too complex. A different approach

could be to use the lake itself as a sentinel for all the processes taking place in and around the lake.

Changes in the surface water solutes can be seen as an integrated response to both catchment specific in-

and out-puts as well as in-lake processes (Krabbenhoft and Webster, 1995; Gurrieri and Furniss 2004;

Rimmer et al., 2006). Using this approach would still call for a somewhat precise estimation of the overall in-

and out-puts of water, but part of the evaluation of the accuracy of the results could be done by simply

studying the surface water. This approach is used in paper 3, and using this method, it was possible to

accurately describe the historical development in alkalinity in the lake over a period of ~40 years.

Even if it was possible to do a truly adequate spatial mapping and quantification of the movement of

water and nutrients in the system, the problems arising from temporal variation would still complicate the

quality of the results. Especially transient effects can be difficult to capture as it is speculated is the case for

Lake Hampen in paper 4. Based on previous studies on the lake (Kidmose et al., 2011; Ommen et al., 2012;

karan et al., 2014) a 200 m2 discharge area was expected to be present at the north-eastern side of the lake.

Instead, this area turned out to be an area recharging the aquifer due to an on-going flow reversal.

Flow reversals are interesting from a hydrological perspective, but also from an ecohydrological

perspective. In lakes where the groundwater normally discharges nutrients to the lake system, a flow

reversal can have significant effect on the system. In the littoral zone, where most of the groundwater under

normal circumstances discharges (McBride and Pfannkuch, 1975; Lee, 1977; Shaw and Prepas, 1990), the

biota benefits from the groundwater supply of nutrients and inorganic carbon (Hargerthey and Kerfoot, 1998;

Frandsen et al., 2012). In lakes with low surface water nutrient concentrations, a flow reversal can cause loss

of an important nutrient source for the vegetation. In lake management, samples of pore water collected

during a flow reversal can be faulty and give rise to wrong estimates of the catchment specific input of

nutrients to the lake. Hence, it is important to be able to both locate and quantify flow reversals.

Groundwater and lake system are connected with each other in a complex manner both temporally,

spatially and disciplinary. It is impossible to accurately describe the hydrology, without taking the biology into

account (i.e. paper 2). On the other hand, it is also impossible to accurately describe the parameters

PhD summary

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controlling growth of the biota without taking the hydrology into account (Frandsen et al., 2012). By

combining methodology from biology/ecology and hydrology, it is possible to produce more accurate

descriptions of the systems, and it often calls for less dense instrumentation of the site of interest. This study

strongly encompasses the idea of ecohydrology as an important new discipline within water management

and sciences.

We live in a world with an ever increasing human population, need for food production, for drinking

water, need for urbanization of our land, and this have wide-ranging consequences for our ecosystems,

including the fresh waters (Moss, 1999). In Demark, we can directly correlate the pollution of our drinking

water with the agricultural use of fertilizers (Hansen et al., 2012), and in countries without regulation of this,

the situation is even worse (Adams, 2001). The need for cross-disciplinarity has never been more urgent.

Furthermore, the idea of cross-disciplinarity might be important on an even higher level than simple

collaboration between different areas of natural sciences. Ultimately, science cannot by itself govern water

management as stated by Nuttle (2002). Public institutions, law, policies and regulation comprise the

material for fashioning sustainable management. At this point however, it can be argued that decision

making e.g. from a cost benefit approach, investment and policy strategies are still not based on a holistic

framework for integrating hydrology and ecology (Zalewski, 2000).

Future decisions concerning our environment are not going to be decided by natural scientists, but

by politicians and economists, consequently natural scientists need to build and strengthen the connection

and collaboration with these and to disseminate the knowledge needed to make the right decisions. As

natural scientists, we must be able to pin-point the challenges posed to our environment, not only within our

own discipline, but between all the disciplines concerning the ecology that we try to protect. We need to

come to a consensus of the actions needed on the political arena and we need to be able to work together

with these people if we hope to have an impact on the decisions made about our ecosystems.

4 References

Adams, W M. 2001. Green development. Environment and sustainability in the third world. Routledge,

London. ISBN (e-Reader format): 0-203-74913-8

Anderson, M. P., Cheng, X. 1993. Long- and short-term transience in a groundwater/lake system in

Wisconsin, USA. Journal of Hydrology. 145(1-2): 1-18

Appelo, C.A. J. & Posta, D. Geochemistry, Groundwater and pollution. 2005. A.A. Balkema publishers. 458-

462

Barko, J. W., Gunnison, D., Carpenter, S. R. 1991. Sediment interactions with submerged macrophyte

growth and community dynamics. Aquatic botany. 41: 41-65

Blais, J. M., Kalff, J. 1995. The influence of lake morphometry on sediment focusing patterns. Limnology and

Oceanography. 40: 582-588

PhD summary

16

Blicher-Mathiasen, G., Rasmussen, A., Grant, R., Jensen, P. G., Hansen, B., Thorling, L. 2013.

Landovervågningsoplande 2012. NOVANA. (Danish Agricultural monitoring catchments). University

of Aarhus. Danish Centre for Environment and Energy, DCE.

Bram, M., Christensen, I. 2006. Management og restaurering af søer i Danmark (management and

restoration og lakes in Denmark). Nordjyllands Amt. Only Danish.

http://naturstyrelsen.dk/media/nst/72307/Management%20og%20restaurering%20af%20s%C3%B8e

r%20i%20Danmark_2006.pdf

Brinson, M. M., Malvarez, A. I. 2002. Temperate freshwater wetlands: types, status, and threats.

Environmental Conservation. 29/2: 115-133

Brock, T.D., Lee, D.R., Janes, D. J. & Winek, D. 1982. Groundwater seepage as a nutrient source to a

drainage lake, Lake Medota, Wisconsin. Water research. 16: 1255-1263

Brock, W. A., Starrett, D. 2003. Managing systems with non-convex positive feedback. Environmental and

Resource Economics. 26: 575-602

Brunke, M. 1999. Colmation and depth filtration within streambeds: retention of particles in hyperheic

interstices. International Review of hydrobiology. Hydrobiology. 84: 99-117

Brunke, M., Gonser, T. 1997. The ecological significance of exchange processes between rivers and ground-

water. Freshwater Biology. 37: 1-33

Brønmark, C., Hansson, L. 2002. Environmental issues in lakes and ponds. Current state and perspectives.

Environmental Conservation. 29/3: 290-306

Carignan, P., Kalff, J.1980. Phosphorus as a source for aquatic weeds: water or sediment. Science. 207:

987-989

Carpenter, S. R., Lodge, D, M. 1986. Effects of submerged macrophytes on ecosystem processes. Aquatic

botany. 26: 341-370

Cheng, X., Anderson, M. P. 1994. Simulating the influence of lake position on groundwater fluxes. Water

Resources Research. 30(7): 2041-2049

Christinsen, P. B., Sørensen, J. 1986. Temporal variation of denitrification activity in plant-covered littoral

sediment from Lake Hampen, Denmark. Applied Environmental Microbiology. 1174-1179

CLEAR –Centre for Lake Restoration. 2006. http://www.lake-restoration.net/english/index_UK.html

Cullmann, J., Junk, W. J., Weber, G., Schmitz, G. H. W. 2005. The impact of seepage influx on cation

content of a central Amazonian floodplain. Jounal of Hydrology. 328: 297-305

Danish Nature Angency. 2014. Basisanalyse for vandområdeplaner 2015-2031 (basis analysis for water area

plans 2015-2031). Danish Nature Agency, Danish ministry of Environment. Available at:

http://naturstyrelsen.dk/media/nst/89701/bilag%201%20basisanalyse%2019-2.pdf

Davison, W. 1986 Internal element cycles affecting the long-term alkalinity status of lakes: implications for

lake restoration. Scheiz. Z. Hydrol. 49/2: 186-201

EEA, 2005. Source apportionment off nitrogen and phosphorus input into the aquatic environment. European

Environmental Agency, EEA, report no.7

Frandsen, M., Nilsson, B., Engesgaard, P., Pedersen, O. 2012. Groundwater seepage stimulates the growth

of aquatic macrophytes. Freshwater Biology. 57/5: 907-92

http://naturstyrelsen.dk/media/nst/89701/bilag%201%20basisanalyse%2019-2.pdf

PhD summary

17

Groffmann, P. M., Pace, M. L. 1998. Synthesis: what kind of a discipline is this anyhow? Successes,

Limintation, and Frontiers in Ecosystem Science. Springer, New York, USA. 473-481

Guildford, J. S. & Hecky, R. E. 2000. Total nitrogen, total phosphorus, and nutrient limitation in lakes and

oceans: Is there a common relationship? Limnology and Oceanography. 45: 1213 – 1223

Gurrieri, J. T., Furniss, G. 2004. Estimation of groundwater exchange in alpine lakes using non-steady mass-

balance methods. Journal of Hydrology. 297: 187-208

Hagerthey, S. E. & Kerfoot, W. C. 1998. Groundwater flow influences the biomass and nutrient ratios of

epibenthic algae in a north temperate seepage lake. Limnology and Oceanography. 43,´: 1227-1242

Hannah, D, M., Wood, P. J., Sadler, J. S. 2004. Ecohydrology and hydroecology: A “new paradigm”?. Hydrol.

Process. 18: 3439-3445

Hansen, B., Thorling, L., Sørensen, B., Dalgaard, T., Erlandsen, M. 2012. Nitrattrends I dansk grundvand.

Vand & Jord. 19 (1): 37-40. Only Danish

Hilton, J. 1985. A conceptual framework for predicting the occurrence of sediment focusing and sediment

redistribution in small lakes. Limnology and Oceanography. 30: 1131-1143

Huettel, M., Gust, G. 1992. Solute release mechanisms from confined sediment cores in stired benthic

chambers and flume flows. Marine Ecology Progress series. 82: 187-197

Huettel, M., Ziebis, W., Forster, S. 1996. Flow-induced uptake of particulate matter in permeable sediments.

Limnology and Oceanography. 41/2: 309-322

Hutchinson G. E. (1975). A treatise on limnology. Limnological botany. John Wiley & son.

Karan, S., Kidmose, J., Engesgaard, P., Nilsson, B., Frandsen, M., Omen, D. A .O., Flint, M. R., Andersen,

F. Ø., Pedersen, O. 2014. Role of a groundwater-Lake interface in Controling seepage of water and

Nitrate. Journal of Hydrology. 517: 791-802

Kidmose, J., Engesgaard, P., Nilsson, B., Laier, T. and Looms, M. C. 2011. Spatial distribution of seepage at

a flow-through lake: Lake Hampen, eastern Denmark. Vadose Zone Journal. 10/1: 110-124

Krabenhoft, D. P., Anderson, M. P., Bowser, C. J. 1990. Estimating groundwater exchange with lakes, 2.

Calibration of a three- dimensional solute transport model to a stable isotope plume. Water

Resources research. 26: 2455 – 2462

Kundzewicz, Z. W. 2002. Ecohydrology – seeking consensus on interpretation of the notion. Hydrological

Sciences journal. 47(5), 799-804

Lee, D. R. 1977. A device for measuring seepage flux in lakes and estuaries. Limnology and Oceanography.

22: 140-147

Lehman, J. T. 1975. Reconstructing the rate of accumulation of lake sediment: The effect of sediment

focusing. Quaternary Research. 5: 541-550

Liboriussen, L., Søndergaard, M., Jeppesen, E. 1999. Sørestaurering i Danmark (lake restoration in

Denmark). Danmark miljøundersøgelse, University of Århus. http://www2.dmu.dk/pub/fr636_del2.pdf

Madsen, T. V., Olesen, B. & Bagger, J. 2002. Carbon acquisition and carbon dynamics by aquatic isoetids.

Aquatic Botany. 73: 371-361

McBride, M.S., Pfannkuch, H. O. 1975. The distribution of seepage within lakebeds. J. Res. USGS. 3: 505-

512

PhD summary

18

Moss, B. 1999. Ecological challenges for lake management. Hydrobiologia. 395(396): 3-11

Nuttle, W. 2002. Is ecohydrology one idea or many? Hydrological sciences journal. 47(5): 805-807

Omen. D. A. O., Kidmose, J., Karan, S., Flint, M. R., Engesgaard, p., Nilsson, B. and Andersen, F. Ø. 2012.

Importance of groundwater and macrophytes for the nutrient balance at oligotrophic Lake Hampen,

Danmark. Ecohydrology. 5/3: 286-296

Rattray, M. R., Howard-Williams, C. & Brown, J.M.A. 1991. Sediment and water as sources of nitrogen and

phosphorus for submerged rooted aquatic macrophytes. Aquatic Botany. 40: 225-237

Rimmer, A., Boger, M., Aota, Y., Kumagai, M. 2006. A lake as a natural integrator of linear processes:

Application to lake Kinneret (Israel) and lake Biwa (Japan). Jounal of Hydrology. 319: 163-175

Rodriguez-Iturbe. 2000. Ecohydrology: A hydrologic perspective of climate-soil-vegetation dynamics. Water

resources research. 36(1): 3-9

Rudd, J. W. M., Kelly, C. A. and Schindler, D. W. 1988. Comment on “Dynamic model of in-lake alkalinity

generation” by L. A. Baker and P.L Brezonik. Water resources research. 24/10: 1825-1827

Sacks, L. A., Herman, J. S., Konikow, L. F., Vela, A. L. 1992. Seasonal dynamics of groundwater lake

interactions at Donãna National Park, Spain. Journal of Hydrology. 135 (1-4): 123-154

Sand-Jensen, K., Prahl, C. 1982. Oxygen exchange with the lacunae and across leaves and roots of the

submerged vascular macrophyte, Lobelia dortmanna L. New Phytologist. 91: 103-120

Schaab, B. F:, Blom-Zandstra, M., Hermans, M. L., Meerburg, B. G., Verhagen, J. 2011. Impact changes of

climatic exstremes on arable farming in the north of the Netherlands. Reg. Environ Change. 11: 731-

741

Sebestyen, S. D., Schneider, R. L. 2004. Seepage patterns porewater, and aquatic plants: hydrological and

biochemical relationships in lakes. Biochemistry. 68: 383-409

Scheffer, M., Hosper, S. H., Meijer, M. L., Moss, B., Jeppesen, E. Alternative equlibria in shallow lakes. Tree.

8 (8): 275-279

Schindler, D. W. 1974. Eutrophication and recovery in experimental lakes: Implications for lake management.

Science. 184: 897-898

Schindler, D. W., Turner, M. A., Staiton, M. P. and Linsey, G. A. 1986. Natural sources of acid neutralizing

capacity in low alkalinity lakes of the Precambrian Shield. Science. 232: 844-847

Shaw, R. D. & Prepas, E. E. 1989. Anomalous, short-term Influx of water into seepage meters. Limnology

and Oceanography. 34: 1343-1351

Sutcliffe, D. W., Carrick, T. R., Heron, J., Rigg, E., Talling, J. F., Woof, C., Lund, J. W G. 1982. Long-term

and seasonal changes in the chemical composition of precipitation and surface waters of lakes and

tarns in the English Lake District. Freshwater Biology. 12(5): 451-506

Søndergaard, M., Jeppesen, E,. Jensen, J. P. 1999. Danske søer og deres restaurering. Miljø og

Energiministeriet. Danmark miljø undersøgelse. Only in danish

Timms, R. M., Moss, B. 1984. Prevention of growth of potentially dense phytoplankton populations by

zooplankton grazing, in the presence of zooplanktivorous fish, in a shallow wetland ecosystem.

Limnology and Oceanography. 29 (3): 472-486

PhD summary

19

Wiberg-Larsen, P., Windolf, J., Bøgestrand, J., Baastrup-Pedersen, A., Kristensen, E. A., Larsen, S. E.,

Thodsen, H., Ovesen, N. B., Bjerring, R., Kronevang, B., Kjeldgaard, A. 2013. Vandløb 2012,

NOVANA (Streams 2012, NOVANA). University of Aarhus. Danish Centre for Environment and

Energy, DCE. http://dce2.au.dk/pub/SR75.pdf

Winter, T. C. 1999. Relation of stream, lakes, and wetlands to groundwater flow systems. Hydrogeology

journal. 7(1): 28-45

Zalewki, M. 2000. Ecohydrology – the scientific background to use ecosystem properties as management

tools toward sustainability of water resources. Ecological engineering. 16: 1-8

Zalewski, M. 2002. Ecohydrology – the use of ecological and hydrological processes for sustainable

management of water resources. Hydrological Sciences. 47(5): 823 – 832

Zalewski, M., Janauer, G. A., Jolankai, G. 1997. Ecohydrology: A new paradigm for the susteineble use of

aquatic resources. UNESCO, Paris. Available at:

http://unesdoc.unesco.org/images/0010/001062/106296E.pdf

http://unesdoc.unesco.org/images/0010/001062/106296E.pdf

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Appendixes

Paper 1

Groundwater Seepage stimulates the growth of

aquatic macrophytes

Mette Cristine Schou Frandsen, Bertel Nilsson, Peter Engesgaard, Ole Pedersen

Published in Freswater biology 2012

Paper 2

Rooted underwater vegetation locally reduces

groundwater discharge in lakes

Mette Cristine Schou Frandsen, Peter Engesgaard, Bertel Nilsson, Ole Pedersen

Paper 2

1

Rooted underwater vegetation locally reduces groundwater

discharge in lakes

Mette Frandsen*, ^

, Peter Engesgaard+, Bertel Nilsson

^ Ole Pedersen

*

*The Freshwater Biological Laboratory, Department of Biology, University of Copenhagen, Copenhagen,

Denmark

^The Geological Survey of Denmark and Greenland, Copenhagen, Denmark

+Department of Geosciences and Natural Resource Management, University of Copenhagen, Denmark

Abstract

A comparative study of the effect of dense stands of Littorella uniflora on the hydraulic properties of the

sediment in Lake Hampen, Denmark was examined on two neighboring areas; one densely vegetated and

one bare. Both areas were located on a groundwater discharge site.

By measuring bulk hydraulic conductivity (Kslichter), vertical hydraulic conductivity (VHC), and doing

analysis of the organic content of the sediment, we found evidence for a strong vegetation induced effect on

the hydraulic conductivity in the lake sediment.

We found an overall negative correlation between Kslichter and the organic content in the sediment

(P<0.001, R2

= 0.47) and a positive correlation between organic content and mass of fine particles (≤63µm)

(P<0.05, R2=0.36). Both Kslichter and VHC was significantly lower on the vegetated sediment compared to the

non-vegetated sediment (P<0.05). On the vegetated area, there was a strong negative correlation between

Kslichter and the mass of fine (P<0.05, R2=0.89) and a strong positive correlation between the mass of fine

particles and the organic content (P<0.05, R=0.89). On the bare sediment the hydraulic conductivity was

relatively uniform with depth compared to the vegetated sediment and there was no correlation between fine

particles, and hydraulic conductivity.

VHC was measured before and after simple disturbance. On the vegetated sediment, the

disturbance had a strong effect in most of the standpipes increasing the VHC significantly (P<0.05). On the

bare sediment the disturbance had little or no effect on VHC.

Introduction

Understanding the exchange of nutrients between groundwater and lakes are important in catchment and

surface water management. Groundwater can accumulate and discharge large amounts nutrients and land

derived pollutants from catchments to lakes. In some catchments, the groundwater discharge represents a

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main transportation route of solutes from the terrestrial to the aquatic environment. In turn, the exchange of

solutes through groundwater discharge can lead to severe impacts on the trophic status of the surface water

(Hagerthey & Kerfoot, 1998; Sebestyen & Scneider, 2004; Hayashi & Rosenberry, 2005; Kidmose et al.,

2013; Ommen et al., 2012; Meinikmann et al., 2013). How groundwater impacts the ecology of lakes varies

from catchment to catchment and depends on the chemical composition of the groundwater and how

important groundwater is to the surface water and chemical budgets.

For groundwater-dominated lakes it is important to understand how the littoral zone of a vegetated

lake bed affects nutrient input and retention (Karan et al., 2014). In the last decades only few studies on lake

ecology includes small to large-scale hydrodynamics of the groundwater-lake interface (Cherkauer & Nader,

1989; Kishel & Gerla, 2002; Rosenberry et al., 2010; Genereux & Bandopadhyay, 2001) with a few studies

including non-reactive tracers such as Chloride and water stable isotopes (e.g. Schuster et al., 2003) and

reactive tracers (e.g. Schafran & Driscoll, 1993). However, direct measurements of the impact of submerged

macrophytes on groundwater seepage have not been carried out although a few studies have indicated a

relationship between presence of macrophytes and groundwater discharge/recharge (Petticrew & Kalff,

1991; Sebestyen & Sneider, 2004; Hayashi & Rosenberry, 2005; Frandsen et. al., 2012; Ommen et al.,

2012; Karan et al., 2014).

It is generally accepted that groundwater discharge in homogeneous aquifers mainly takes place

near the shore and decreases exponentially with distance to shore (McBride & Pfannkuch, 1975; Lee, 1977;

Shaw & Prepas, 1990). In many cases, however, sediment-dependent deviations from this exponentially

decreasing pattern have been observed. In non-homogeneous aquifers the local discharge patterns strongly

depends on small scale (cm to meters) hydraulic properties of the lake bed sediments. For example, Kishel &

Gerla (2002) and Sebok et al. (2013) found small-scale irregular discharge patterns controlled by the

heterogeneity of the lake bed and aquifer. Thus, an interesting question is what controls the hydraulic

properties of lake beds? Some of the more well-known parameters are those that are of strictly

physical/geological (Benoy & Kalff, 1999; Rosenberry et al. 2010) or chemical nature (Schälchli, 1992;

Bouwer 2002; Förstner et al. 2008), whereas biological effects have been less studied (Schälchli, 1992;

Brunke, 1999; Marmonier et al., 2004). Physical effects can be very local. For example, Rosenberry et al.

(2010) used seepage meters to examine recharge (surface water lost to groundwater) before and after

simple disturbance of the top sediment (here by just walking on the sediment) and found a 2-7 fold increase

in seepage rates after disturbing the top layer. They attributed this to the existence of a veneer layer on the

lake bed.

An important, but overlooked biological effect is the one caused by the roots of submerged plants.

The roots provide habitat for sediment biota (i.e. microbes and invertebrates) hereby having a direct physical

effect on the sediment. The direct biological/physical effect may in turn have an indirect chemical effect by

affecting the oxygen penetration depth (Forster & Graf, 1995; Michaud et al. 2005). Furthermore, the

presence of the roots themselves also affects the porosity of the sediment (Palmer et al., 2000) but it is

unknown how this works in lake settings. Moreover, important inter-dependent interactions between physical

and biological processes are the presence of a canopy above the sediment (i.e. the leaves). For example,

physical colmation primarily happens in the littoral zone that supports large stands of submerged

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macrophytes. Here, submerged plants attenuate the energy associated with waves and currents creating a

low energy environment that allows entrapment of finer sediments (Hilton, 1985; Blais & Kalff, 1995). The

intrusion of fine particles (≤63 µm) (Lehman, 1975; Hilton, 1985) reduces sediment porosity and may cause a

significant decrease in hydraulic conductivity. This is a well-known phenomenon in streams (Brunke &

Gonser, 1997; Huettel et. al., 1996; Huettel & Gust, 1992), but has not been examined thoroughly for lake

settings.

Lakes represent low energy environments compared to streams. However, over longer time scales

the physical and plant related processes in the littoral zones may act in the same way as in flowing streams

by stabilizing the near shore areas from corrosion by attenuating the energy from current and waves. In

areas of high plant abundance these processes may effectively decrease the porosity and permeability of the

lake bed covered with plants.

Groundwater discharging near the lake shore percolates the littoral sediments where it surpasses

the 10- to 15-cm thick rhizosphere before entering the lake. Rooted aquatic plants such as isoetids (Robe &

Griffiths, 1998; Andersen & Andersen, 2006; Pedersen et al., 1995; Sand-Jensen et al., 2005) as well as

other biota associated with the sediment-water interface (SWI) filter nutrients (Loeb & Hackley, 1988; Wetzel

1990; Hagerthy & Kerfot, 1998; Frandsen et al. 2012) and inorganic carbon (Sand-Jensen et al., 1982,

Madsen et al., 2002; Winkel & Borum, 2009) carried with the groundwater explaining why only a fraction of

the nutrients present in groundwater ever enters the lake (Ommen et al. 2012). Groundwater not surpassing

this littoral filter, but discharging at greater depth beyond the littoral zone, may enter the lake relatively

unfiltered compared to the discharge in the littoral zone.

The filtering capacity of the submerged macrophytes varies between species. Isoetids such as

Littorella uniflora have a very efficient root uptake as their roots often comprise up to 50% of the total plant

mass. In contrast the elodeids have very small roots compared to the shoots and lack the internal lacunae for

effective internal transport of nutrients. Accordingly, these species react differently to groundwater seepage.

This was demonstrated by Frandsen et al. (2012) who showed that the input of nutrients and inorganic

carbon with the groundwater seepage significantly affected the growth of the L. uniflora, whereas the elodeid

Myriophyllum alternifolium showed little to no response.

What also makes the isoetids more interesting when it comes to evaluating the interaction between

aquatic macrophytes, seepage, and nutrient conditions in lake beds is the community structure. Some

species of isoetids tend to form carpet-like turfs holding up to 9.000 macrophytes per m2 (Christensen &

Sørensen, 1986). With the large root systems of isoetids, this may effectively alter the hydraulic properties in

the root zone of the sediment.

The existence of a vegetation-induced lowering of the hydraulic conductivity in lake bed settings

have been suggested a couple of times in the literature. Karan et al. (2014) found evidence of lowered

hydraulic conductivity in the littoral zone, but based on a modelling approach. They identified an off-shore

discharge peak, after the near-shore vegetation cover, and could only explain this by including a lower-

permeable lake bed near the shore forcing the model to simulate groundwater to discharge further off-shore.

This, however, was only indirectly linked to the vegetation cover. Hargerthey & Kerfoot (1998) found a

several fold increase in discharge when installing custom-made seepage growth chambers consisting in the

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lake bed. Similar observations were done by Frandsen et al. (2012) who found a several fold increase in

discharge after installing similar growth chambers. They speculate if the installation of the growth chambers

might have punctured a layer of low hydraulic conductivity.

The existence of a vegetation-induced lowering of the hydraulic conductivity is interesting in many

ways. First of all, it would lead to an increased discharge further away from the shore to off-shore areas with

no effective plant cover filtering the water. From a lake management point of view diverted groundwater

seepage could represent an undetected source of nutrients. Also, when modelling groundwater-lake

interactions, this could be an important factor. While some studies (e.g. Kidmose et al., 2011, 2013) have

included lower-permeability areas in the off-shore regions caused by the presence of sedimentary organic

material, it could be assumed that plant covers in the littoral zone could have an equally important effect.

In this study we examine the relationship between hydraulic conductivity and small-scale

sedimentary conditions caused by the presence of vegetation by comparing results from both a vegetated

and a non-vegetated area of a lake bed in a groundwater-fed lake. We hypothesize that the hydraulic

conductivity is affected by the vegetation, leading to a reduced hydraulic conductivity in area with dense

plant cover. We further hypothesize that this is an effect of colmation of fine particles in the vegetated

sediment.

Materials and methods

Study site

Lake Hampen (surface area 76 ha; maximum depth 13 m; mean depth 4 m) is a Northern temperate lake,

located in the western part of Denmark just east of the last glacial advance (Figure 1A).

The lake is situated in a 15-25 m deep layer of coarse melt water sands and gravel. The lake is

characterized as a flow-through lake, with the groundwater flowing in a North-East, East and South-East to

West direction. Approximately 2/3 of the water is received from groundwater discharge and 3/4 of the water

leaves the lake through groundwater recharge at the western shore line. There are no major in- or outlets.

Precipitation and evaporation account for the remaining in and out fluxes (Ommen et al., 2012; Kidmose et

al. 2011). Lake Hampen is an oligo- to mesotrophic softwater lake with an annual mean alkalinity of 0.15

mM. The mean summer concentrations of total Phosphorous (TP), total Nitrogen (TN) and planktonic

chlorophylla are 0.05, 0.16, and 0.005 mg L-1

, respectively (mean summer values 1971 to 1999, Moeslund,

2000). The lake catchment (993 ha) is primarily covered with forest (62%) and agricultural land (30%). The

agricultural land is located near the North-Eastern shoreline, near the main discharge area which is the main

focal point of the present study (Figure 1B).

To test the hypothesis that rooted submerged vegetation could affect the hydraulic properties of

lake beds, we performed an in situ comparative study between two areas established on the North-Eastern

shoreline; one with a dense plant cover of Littorella uniflora (vegetated area) and one without vegetation

(non-vegetated area) (Figure 1B). The vegetated area was right next to the farmland and the non-vegetated

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area was located 300 m to the North-West. The bulk and vertical hydraulic conductivities in both areas were

estimated based on grain size distributions of lake bed sediments and falling head experiments, respectively.

The organic content in the lake bed sediments was measured and all data was used to perform simple

correlation analyses to investigate possible relationships between plant cover (as reflected by organic

content), colmation (as reflected by grain size distribution), and hydraulic conductivity.

Instrumentation

Vertical hydraulic conductivity

In situ vertical hydraulic conductivity (VHC) was measured using a standpipe method (Chen, 2000). Rigid

Plexiglas pipes (diameter = 5 cm) were pushed vertically down (~30 cm) in the lake bed sediment (Figure 2).

The pipes were filled with water and the head drop per unit time was measured. The VHC was calculated

using;

( ) (

) (1)

Figure 1) Map of Lake Hampen (A) with zoom in on North-Eastern site of the lake showing

conceptual map of instrumentation of the vegetated and the non-vegetated area (B). Grey dots

illustrate standpipes and grey areas illustrate vegetated sediment. The distance between the

vegetated and the non-vegetated area was approximately 300 m.

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where Lv = the height of the sediment core inside the pipe; h1 and h2 are the hydraulic heads at time t1 and t2,

respectively. Equation (1) does not account for anisotropy, but Chen (2000) demonstrated that as long as the

ratio of the sediment column (Lv) and the diameter of the the pipe (D) is larger than ~3 (Lv/D > 3) the error by

using (1) is small for realistic anisotropy ratios (10). Sebok et al. (2014) also found that anisotropy had

limited influence on the estimation of VHC and that Equation (1) worked well for stream bed sediments.

Standpipes were installed perpendicular to the shore ~4-18 m from the shoreline. On the non-

vegetated area 7 standpipes were installed. It should be noted that the standpipe furthest from shore, was

installed on vegetated sediment (Standpipe No. 7).

On the vegetated area 12 standpipes were installed

(Figure 1B)

The experiments were carried out two

times. In August 2013 the VHC was estimated on

both areas as described above, i.e. with a falling

head experiment on an undisturbed SWI. The

second time, we disturbed the top sediment in the

VHC pipes before starting the experiment. The

sediment was disturbed by punctuating the top 10

cm of the sediment inside the standpipes with at

spear (diameter = 1-5 cm) approximately 10 times in

an attempt to destroy a veneer or colmation layer.

Subsequently, the falling head measurements were

repeated as described above. One of the stand

pipes in the non-vegetated area was lost during the

disturbance experiment, and is thus not represented

in the comparative experiment. The mean VHC for

the two areas (vegetated and non-vegetated) were

compared using a Mann-Whitney U-test for non-

Gaussian distributed data. The analysis was

performed on mean VHC values for each standpipe

using a significance level of α = 0.05. Differences in

VHC before and after the disturbance were tested using at t-test for paired data using a significance level of

α = 0.05. Variance homogeneity for all the VHC estimates for each standpipe was tested with an F-test.

Grain size analysis

Analysis of the grain size distribution was made on sediment cores from both vegetated and non-vegetated

sediments collected in May 2011. The sediment cores were sampled randomly by pushing small Plexiglas

tubes vertically into the sediment (N = 4, D = 5 cm, length = 14-16 cm). The cores were frozen and

subsequently cut in half along the vertical axis. Each half core was cut in 2 cm slices, resulting in a total

Figure 2. Schematic diagram of a standpipe for measuring the vertical hydraulic conductivity. The standpipe was pushed vertically into the lake bed sediment (Lv). The difference between the hydraulic head in the stand pipe and the lake stage induced a flow through the column and the change in water column level in the pipe (h1-h2 etc.) can be used to calculate the vertical hydraulic conductivity.

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number of samples for grain size analysis of 60-62 from each area. Sieving of each layer was performed on

dry material using meshes of 1000, 710, 500, 355, 180, 150, 125, 63, 20 and 10 µm. Particles smaller than

125 µm were measured by a laser diffraction particle size analyzer (ISO 13320,Mastersizer, Malvern, UK).

Bulk hydraulic conductivity (K) was calculated using the Slichter formula as suggested by Vukovic & Soro

(1992);

(2)

where g is gravitational acceleration of 9.82 ms-2

, is kinematic viscosity of 1.14*10-6

m2 s

-1 (at 16°C), Cs is a

unit less coefficient of 1*10-2

, n is porosity empirically estimated from n = 0.255 * (1+0.83Cu

), where Cu =

d60/d10 is the coefficient of uniformity, d10 and d60 are the grain diameter (m) at which 10 and 60 % of the

sediment is finer, respectively. In this study, however, we used a constant porosity constant of 0.32

representative for sandy sediments. The Slichter formula is appropriate for samples with grain diameters in

the range 0.01 to 0.5 mm.

To compare the bulk hydraulic conductivity with VHC, we calculated the harmonic mean (Fitts,

2013). The harmonic mean takes into account that different layers of sediment have different hydraulic

conductivity, and gives a weighted average hydraulic conductivity across all the layers. The harmonic mean

was calculated as VHCg = ∑di / ∑(di/Ki), where VHCg is the vertical hydraulic conductivity of the core

estimated from grain size distribution, di is the thickness of the ith layer, and Ki is the bulk hydraulic

conductivity of the ith layer.

The mass of fine particles was calculated as mass of fine particles ≤63 µm per total mass of

sample. Differences in mass of fine particles between the two areas were calculated using at t-test for

unpaired data at a significance level of α = 0.05.

Organic content in the sediment was used as a proxy for plant density and measured in each of the

2 cm thick sediment slices. Even though organic content is not strictly related only to plant density as it

integrates all the organic content in the sediment including debris, animals and others, it can be used in a

comparative way as we had both a vegetated and an non-vegetated transect. The sediment was dried at 105

°C for a minimum of 24 hours and organic matter was measured as weight loss on ignition (550°C).

Differences in organic content between the two areas were calculated using at t-test for unpaired data using

a significance level of α = 0.05. Correlation analysis was carried out between the mass of fine particles, the

organic matter, and the hydraulic conductivity using Pearson correlation analysis for Gaussian distributed

data (if not Gaussian distributed a spearman correlation analysis was used) using a significance level of α =

0.05.

Results

Vertical hydraulic conductivity - Stand pipe experiments

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There were significant differences in VHC between the two areas in the first set of experiments (no

disturbance). The VHC on the vegetated area was significantly lower compared to the non-vegetated area (P

< 0.05) (Figure 3). The mean VHC of the non-vegetated area (5.4 md-1

, N= 7) is almost four times higher

compared to the vegetated area (1.0 m d-1

, N= 12).

When comparing the overall average (all standpipes within an area) before and after the

disturbance experiment, no significant differences were found on either of the areas (P > 0.05). However, by

inspecting the VHC before and after the disturbance in the individual stand pipes, the reason for this is clear.

On the vegetated area, the VHC was significantly higher after disturbance compared to before disturbance in

9 of the 12 standpipes, whereas the remaining 3 standpipes exhibited the opposite behaviour (Figure 4). In 4

of these 9 stand pipes (a1, a2 , a5 ,a7) the VHC was not detectable in the experiment before disturbance.

After disturbance, VHC in these four standpipes was measured to approximately 1.8, 0.9, 0.8, 0.9 m d-1

,

respectively, which is close to the overall average VHC after disturbance of 1.1 m d-1

. On the non-vegetated

sediment, the differences in VHC before and after the disturbance were only significant in 2 of the 7

standpipes (P < 0.05) (standpipe 1 and 6, Figure 5). Here, VHC increased from 2.9 to 3.9 m d-1

(standpipe 1)

and from 1.5 to 3.1 m d-1

(standpipe 6).

Bulk hydraulic conductivity and organic matter

There were significant differences in the distribution of bulk hydraulic conductivity (based on grain-size

distribution) between the two areas (P < 0.05). The hydraulic conductivity was more than 5-fold higher in the

non-vegetated areas (mean ± sd = 8.8 ± 1.7 m d-1

, number of cores N = 4, number of samples n = 62)

compared to the vegetated sediment (mean = 1.6 ± 1.0 m d-1

; N= 4, n = 62) (Figure 6). On the vegetated

area the hydraulic conductivity decreased from 4 to

8 cm depth, followed again by an increase below

this point. The lowest hydraulic conductivity was

found at depths 6 to 8 cm ( 0.5 m d-1

). The highest

hydraulic conductivity ( 4 m d-1

) was at 14 cm depth

(Figure 7). The above pattern was not seen on the

non-vegetated area. Here, the hydraulic conductivity

decreased from 12 m d-1

to 6 m d-1

from the top of

the sediment and down to 8 cm depth. Below 8 cm

depth, the hydraulic conductivity was nearly

constant (Figure 7).

There were significant differences between

the mass of fine particles in the sediments of the two

areas (P < 0.05, Figure 8), and two different

distributions patterns occurred with depth on the two

areas (Figure 9). On the vegetated area, the mass

of fine particles followed an increasing pattern from

the top of the sediment and down to 8 cm. This was

Figure 3. Box-whiskers plot showing the difference in vertical hydraulic conductivity (VHC) on the vegetated area and the non-vegetated area. The plot shows median (horizontal line), upper and lower quartiles (boundary of the box) and the whiskers show minimum and maximum values. The difference in vertical hydraulic conductivity between the two areas was significant (P < 0.001; Mann Whitney test).

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9

also reflected in the D10, which followed an opposite pattern than that of the mass of fine particles and a

similar pattern as the hydraulic conductivity (Figure 7, 9 and 10). Interestingly, both the D10 and the

distribution of fine particles on the non-vegetated sediment were almost the same in every sediment layer

(Figure 9, 10)

An overall strong negative correlation was found between organic matter and average bulk

hydraulic conductivity (P < 0.001; R2

= 0.47) (Figure 11a). In agreement with this, a strong positive

correlation between organic matter and the mass of fine particles was found (P < 0.005; R2

= 0.36) (Figure

11b). Looking at the two areas separately, some notable differences were observed. The organic content in

the sediment on the vegetated area was significantly higher compared to the bare sediment (P < 0.05)

(Figure 8). On the vegetated area, the organic content was positively correlated to the mass of fine particles

(P < 0.05; R = 0.89), and negatively correlated to the hydraulic conductivity (P < 0.05; R = 0.89). On the non-

vegetated area, the organic content was more or less the same in all layers of the sediment and it was 5 to

Figure 4) Vertical hydraulic conductivity (VHC) before (U) and after (D) disturbance on the vegetated area.

The vertical hydraulic conductivity was significantly higher after disturbance compared to before disturbance

in all except three standpipes. In four standpipes (a1, a2, a5, a7), the VHC went from being un-detectable to

relatively high rates as caused by the disturbance. In standpipe 4, 5 and a4, the VHC was significantly lower

after the disturbance. * means P < 0.05; and ** P < 0.005 (t-test for matched pairs).

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10-fold lower compared to the vegetated sediment (Figure 8, 9). No correlation was found between the

organic content and hydraulic conductivity and the mass of fine particles (P > 0.05).

The coefficient of uniformity (Cu) has a big impact on bulk hydraulic conductivity (Equation 1). Cu

was in general slightly higher on the vegetated area compared to the non-vegetated area revealing that the

grading of the sediment particles in the vegetated area was much higher compared to the non-vegetated

area (Figure 7). In the vegetated sediments, the maximum Cu was reached at 8 cm depth and the distribution

followed an almost opposite pattern to that of the hydraulic conductivity but similar to that of organic content

and mass of fine particles (Figure 7 and 9). A very steep increase in Cu was observed at the same depth,

where a high mass of fine particles and low hydraulic conductivity was found (Figure 7 and 10). In the non-

vegetated sediments Cu was highest near the top of the sediment and decreasing from 2-8 cm depth

followed by a stable level from 8-16 cm. In contrast, the Cu followed the same pattern as that of the hydraulic

conductivity on the non-vegetated area (Figure 7). In the top 6-7 cm, the Cu was high followed by a more or

less constant value in the deeper parts. The high Cu corresponds to a layer with more coarse gravel we

Figure 5). Vertical hydraulic

conductivity (VHC) before (U)

and after (D) disturbance on the

non-vegetated area. There was

no significant effect of the

disturbance in all except two

standpipes. In standpipe 1 and

6 the vertical hydraulic

conductivity significantly

increased after disturbance. *

means P < 0.05; and ** P <

0.005 t-test for matched pairs

Figure 6) Box whiskers plot

showing the difference in bulk

hydraulic conductivity (Kslichter) on

the vegetated and the non-

vegetated area. The plot shows

median (horizontal line), upper

and lower quartiles (boundary of

the box) and the whiskers show

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observed in the top layers of the sediment. The gravel layer can explain why both the hydraulic conductivity

and the Cu have these distinct high values in the top layers of the sediment, whereas the organic content and

the mass of fine particles have uniform values for all depths.

Discussion

This study demonstrates that dense bottom vegetation promotes the development of a layer in the lake bed

with lower hydraulic conductivity. A similar phenomenon has been observed in streams (Brunke & Gonser,

1997; Huettel et al., 1996; Huettel &

Gust, 1992), but not for lakes although a

few studies have speculated on a

similar effect (Karan et al., 2014;

Hagerthy & Kerfoot, 1998; Frandsen et

al., 2012). The reduction in hydraulic

conductivity depends on mainly two

mechanisms; the intrusion of finer

particles (≤63 µm) in the top layers of

the sediment and presence of the

macrophytes themselves as indicated

by the elevated organic content in the

sediments with lower hydraulic

conductivity.

Dense vegetation lowers the

hydraulic conductivity

The comparative study of the VHC both

before and after disturbance shows that

vegetation causes the overall VHC to be

lower on a transect with dense

vegetation cover compared to a

neighboring transect without vegetation

clearly indicating the existence of a

vegetation effect on the hydraulic

conductivity (Figures 3, 4, 5, 7).

On the vegetated site we

found an overall increase in the VHC

after disturbance although the effect

was blurred by the opposite effect in

Figure 7. Bulk hydraulic conductivity (KSlichter) and the

Coefficient of Uniformity (CU = D60 / D10) at different

depths in the sediment in the vegetated and non-

vegetated area. Graphs show mean (solid line) ± SE

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some of he pipes. The VHC increased significantly in all except three standpipes (Figure 4). Furthermore, in

four standpipes the lake bed went from being almost completely clogged (i.e., no water flowed out of the

pipes during the experiment) to being permeable with a VHC corresponding to the average for the whole

area (Figure 4). Interestingly, the opposite effect was observed in three of the standpipes, i.e., the VHC

decreased as an effect of the disturbance (Figure 4). It could be due to mechanical-induced collapse of fine

tunnels in the sediment, keeping the VHC high before the disturbance.

At the non-vegetated area the overall effect of disturbance was very small. Only in two of the

standpipes a significant increase in the VHC was found, namely the pipe closest to the shore (pipe 1) and

one of the pipes furthest from shore (pipe 6) (Figure 5). It should be noted that pipe 6 was placed within a

vegetated zone, and the effect of the disturbance could be related to the plant cover (Figure 1). The effect on

the VHC in pipe 1 could reflect some natural variability in grain size distribution and perhaps a thin clayish or

organic layer in this pipe was punctuated by the disturbance. Still, the overall results show that disturbance in

this area had little or no effect on VHC.

Similar effects of sediment disturbance leading to higher seepage fluxes have been observed by

Rosenberry et al. (2010), where only the top of the sediment was disturbed (simply by walking over it). This

was not, however, associated with a vegetation cover. Frandsen et al. (2012) and Hagerthey & Kerfoot

(1998) both observed increased fluxes after disturbing the sediment by implementing seepage growth

chambers in the sediment. In both cases the increased fluxes were not correlated to the plant cover, but

observed as an effect of removing the top sediment replacing it with technical sand (with a higher porosity

than the natural sediment).

Figure 8) Average of fine

particles and D10 compared

to the relative mass of

organic matter in the

sediment on the vegetated

and the non-vegetated

area. There was a

significant difference

between the mass of fine

particles (P < 0.05, t-test)

between the vegetated and

the non-vegetated area.

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A similar linkage between hydraulic conductivity and plant cover have been established for lotic

systems, where it is explained with the intrusion of fine particles in the low energy environment present in

macrophyte stands (Brunke & Gonser, 1997; Huettel et. al., 1996; Huettel & Gust, 1992). Our results point to

similar effect in lentic systems.

Intrusion of fine particles in the sediment

The grain size analysis explains some of the observed VHC patterns. We found significantly higher relative

mass of fine particles on the vegetated area compared to the non-vegetated area (Figure 8). When

averaging all the estimates of K for each 2 cm layer a distinct pattern in the hydraulic conductivity was found

on the vegetated area. The lowest hydraulic conductivity coincides with the sediment depth of normally

identified with the highest root density for isoetids (6-8 cm) (Møller et al., 2013). This support the evidence for

a vegetation-induced lowering of the hydraulic conductivity as found in the stand pipe experiment.

To explain the differences observed in hydraulic conductivity between the two areas the differences

in Cu and the distribution of fine particles between the two areas were evaluated. At the vegetated area, a

high Cu coincides with a low hydraulic conductivity, whereas the highest Cu coincides with the highest

hydraulic conductivity on the non-vegetated area (Figure 7). This can be explained by the distributions of fine

particles and D10 (Figure 9). In the vegetated area, the high Cu negatively correlates with the mass of fine

particles and D10 (Figure 9). On the non-vegetated sediment, a layer of gravel and fine sand in the top of the

sediment causes the hydraulic conductivity and the Cu to be very high making them positively correlated.

The difference between the two areas in respect to the vertical distribution of fine particles

corresponds well to previous studies (Descloux et al., 2010; Petticrew & Kalff, 1991; Sand-Jensen, 1996)

showing that the low energy environment in the littoral zone induce entrapment of fine sediment particles that

otherwise only accumulates in the profundal zone (Benoy & Kalff, 1999).

Figure 9) Comparison of

organic matter and

relative mass of fine

particles (≤63µm) in the

sediment in the

vegetated and the non-

vegetated area.

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Correlation between organic matter and

mass of fine particles

The strong correlation between organic matter

and mass of fine particles and between

organic matter and hydraulic conductivity

(Figure 11) strongly indicates a connection

between plant cover and hydraulic

conductivity. The above ground biomass

ensures entrapment of the fine particles

whereas the below ground biomass possibly

holds the fine particles in the top part of the

sediment.

Normally the entrapment of fine

particles in macrophyte stands is associated

with the low energy environment in the

canopy of the macrophytes (Lehman 1975;

Hilton, 1985; Blais & Kalff, 1995). Sediment

entrapment of fine sediments has only

recently been investigated in lacustrine settings. Petticrew & Kalff (1991) primarily focus on the differences in

flow rates within macrophytes stand. They find that the flow rates are lowest where the plant surface area to

water column ratio (Ps/Wvol) is lowest, i.e., the energy attenuation is affected by the above ground biomass.

Also, Losee & Wetzel (1993) used a modelling approach to show that entrapment of fine particles primarily

happens in macrophyte stands. In this study we investigated sediment/vegetation interactions in stands of

isoetids reaching a maximum high of ~10 cm. The energy attenuation is somewhat limited compared to

elodeid stands (reaching sometimes meters) due to the low Ps/Wvol ratio. However, these species have very

large root systems compared to elodeids, and based on the results from this study we speculate that

entrapment of the fine particles over time, might be more efficient as the roots perhaps trap and hold the fine

particles in the sediment.

Roots may physically alter the hydraulic properties of the sediment

The presence of roots can affect the hydraulic conductivity in more ways, than by entrapment of fine

particles. Vegetation physically affects the structure of the sediment. The physical effect of plant roots on the

soil structure is a well-studied field in wetland ecology and on dry land surfaces (Rillig & Mummey, 2006;

Gyssels et al. 2005). On land, macrophyte roots create macro pores that favours water bypass increasing

the hydraulic conductivity of the soil (Mann & Wetzel, 2000; Angers & Caron, 1998; Mitchell & Ellsworth,

1995). How the roots physically affect the hydraulic conductivity under submerged conditions have not been

investigated thoroughly. Some studies show that roots act to stabilize sediment against erosion and, as

discussed above, generate an environment that favours accumulation of fine particles (Gregg & Rose 1985;

Iversen et al. 1985). Sand-Jensen (1998) shows a correlation between organic matter and the amount of fine

Figure 10) Comparison of organic matter and D10 in the

sediment of the vegetated and the non-vegetated area.

Paper 2

15

sediment. Studies focusing on these

interactions with focus on hydraulic conductivity

are however non-existent to our knowledge.

It could be argued that the active

roots take up pores spaces otherwise open for

water flow hereby physically decreasing the

porosity of the sediment leading to a lowered

hydraulic conductivity similar to the process of

colmation. L. uniflora has an average root

diameter of ~0.25 mm (Raun et al., 2010) and

average length of 100 mm. They grow in carpet

like structures holding around 9000 individuals

per m-2

(Christensen & Sørensen, 1986).

Assuming that the pants have as a minimum

five leafs and two roots pr. Leaf gives a total

number of 90.000 roots per m2, with a total root

volume of 0.43 10-3

m3. The porosity of this

sediment is approximately 0.35 why the volume

of the pore spaces in the sediment is

somewhere around 0.035 m3; thus the roots

may occupy 1,2% of the total pore space. As

the linkage between hydraulic conductivity and

porosity is nonlinear, this would greatly affect

the hydraulic conductivity, especially because

the roots might occupy the larger pore spaces. For example, assume a cubic packing of the soil grains with a

diameter of 0.5 mm, then the intra-pore volume is equivalent to a pore with a diameter of 0.26 mm or similar

to the root diameters of L. uniflora. Given that the roots occupy the pore spaces already present in the

sediment they could theoretically fill the biggest pore spaces hereby greatly affecting the hydraulic

conductivity.

Despite the strong correlation we found between biomass and mass of fine particles, further

studies must be done to shed light on the connection between hydraulic conductivity and the presence of

plant roots. In this study we did not correlate the number of roots with the hydraulic conductivity, but used

organic content as a proxy for this. The total amount of organic content does not give a true indication of the

root mass present as other types of organic substances are present in the sediment. It could be speculated

that what we see is a secondary effect of the presence of plants. Some of this organic matter is in the form of

debris, and it is possible that it is the decomposed organic matter that creates the effect we see. Still, as we

did a comparative study we are able to detect differences between the vegetated area holding sediment with

large volume of living roots and non-vegetated area holding little or no living roots.

Figure 11) Correlation of organic matter with mass of

fine particles and bulk hydraulic conductivity. There

was a strong correlation between organic matter and

KSlichter (P < 0.001, R2 = 0.47) and between organic

matter and fine particles (P < 0.005, R2 = 0.36)

Paper 2

16

We used two methods to examine the hydraulic conductivity, namely standpipe experiments giving

the in situ vertical hydraulic conductivity (VHC) and grain size analysis giving the bulk hydraulic conductivity

of a disturbed sample (Kslichter). The VHC estimations are lower compared to the Kslichter on both the vegetated

and the non-vegetated area. This is in agreement with Song et al. (2009) who explain the differences by

especially the disturbance of the samples used for grain-size analysis. By doing so the vertical anisotropy is

somewhat removed. Under in situ conditions the sediment will have interbedded layers of coarse and fine

grained layers. The normal sedimentary structure of streambed sediment will have a higher horizontal

hydraulic conductivity than a vertical hydraulic conductivity. This is probably part of the reason for some of

the observed differences between VHS and K observed in this study.

Landon (2001) found that hydraulic conductivity estimated from empirical grain size formulas are

generally lower than those determined from field scale hydraulic tests. This is attributed to differences in

spatial scale in thecompared In situ tests. Landon (2001) further suggest that the hydraulic conductivity

calculated using the grain size distribution may be smaller because K is a complex function of packing,

sediment structure, heterogeneity and other factors not accounted for in the empirical grain size method

(Tayler et al. 1990, Landon et al. 2001). Song et al. (2009) suggest that the Landon findings are due to the

different depth used to measure vertical and bulk hydraulic conductivity respectively. I.e., Landon measured

vertical hydraulic conductivity in deeper sediments than the sediment used to conduct gran-size analysis. In

the same study they found a decreasing vertical hydraulic conductivity with depth using slug test. Song et al.

(2009) also points to the use of the empirical coefficient in the Hazen formula as part of a possible

explanation

Another explanation is given by Sebok et al. 2014 who also suggest that even small amount of

organic matter in the sediment can skew the in situ measurements of hydraulic conductivity compared to bulk

measurements where the organic content is not taken into account. This is in good agreement with our

study. Furthermore using equation (1) a small underestimation of VHC is likely as described by Chen (2000).

Comparing our results to other studies conducted at Lake Hampen there are some differences.

Karan et al. (2014) found the horizontal hydraulic conductivity (Kh) in the aquifer near the vegetated area

using slug tests and grain size analysis. They estimated the Kh to 30 m day-1

and calibrated an anisotropy of

50 in order to fit the observed discharge distribution at the same site as the vegetated area examined in this

study. The calibrated vertical hydraulic conductivity of the aquifer sediments therefore equals 0.6 m day-1

.

The vertical hydraulic conductivity in the aquifer is similar although a factor of about 2-3 lower than the VHC

we estimated for the vegetated area (undisturbed). To compare this with the hydraulic conductivity based on

grain size, the harmonic mean is calculated (Fitts, 2013). The harmonic mean gives a vertical hydraulic

conductivity VHCg = 1.0 m day-1

at the vegetated area. This value is similar to that used in the modelling

study of Karan et al. (2014). However, to better match simulated and observed discharge Karan et al. (2014)

also tested the inclusion of a top 15-cm-thick plant zone with a much lower horizontal hydraulic conductivity

of 0.025 m/d in the first 22 m of the littoral zone. The anisotropy of this zone was also given as 50 implying

that the vertical hydraulic conductivity is as low as 0.0005 m/day. This is much lower than what we measure

using the stand pipes or grain size. This could be caused by several things. First of all, the study by Karan et

al. (2014) is a modelling-based study and as they argue they were not able to find a unique parameter set.

Paper 2

17

For example, they show that the vertical hydraulic conductivity of the plant zone may be as high as 0.025

m/day and still match observations satisfactorily. However, this is still much lower than what we measure.

Secondly, the lake bed in Lake Hampen is very heterogeneous, why a single stand pipe represent perhaps

numerous different layers with different hydraulic conductivity and it is possible that with the 12 stand pipes in

the vegetated area do not capture the true variability in hydraulic conductivity of the vegetated zone. Here it

is worth noting that actually 30% of the stand pipes allowed no flow to begin with (Figure 4). This is a rather

large proportion of the lake bed. Installation of the standpipes can generate boundary effects along the sides

of the standpipe, for example in the form of small rubbles generating preferential flow paths near the walls of

the pipe. When pushing the standpipe down through the sediment, the sediment is perhaps compressed in

other areas. These things would for the most part lead to artificial high fluxes within the standpipes. However,

as we used the same method at both areas and mainly look at the relative differences between the two

transect, this should not affect the overall results when comparing a vegetated and a non-vegetated area.

This can cause problems when comparing the relative magnitude with those obtained by other methods.

It seems likely, that the carpet like structures of L. uniflora plays a complex role in the lake ecology.

In this study we find evidence that dense carpet structures of L. uniflora affects the groundwater seepage

discharge by lowering the hydraulic conductivity in the top layer of the sediment hereby diverting the

subsurface groundwater flow to discharge further off shore where the vegetation is sparse. As we have

shown previously groundwater seepage discharge positively affect vegetation growth (Frandsen et al., 2012).

The vegetation represent a filtering layer of the incoming groundwater, why a diverted flow leading to

discharge of water in non-vegetated zones, could impact the surface water ecology as well. Finally, at Lake

Hampen the groundwater discharge zone coincides with a nearby agricultural farmland and large amount of

nitrate enters the lake through the groundwater (Ommen et al., 2012). This groundwater contains nitrate

concentration reaching 150mg L-1

. If the groundwater enters the lake beyond the vegetated zone this could

enter the lake relatively unfiltered. It could be an interesting study to examine how important the vegetation of

isoetids are for protecting lakes like Lake Hampen from nutrient enrichment through groundwater discharge.

Acknowledgement

We would like to give a special thanks to Jens Bisgaard (GEUS) for creative and technical

assistance. A warm thank to Mikkel René Andersen for valuable help in the field. We acknowledge the kind

support of the Centre for Lake Restoration, a Villum Kann Rasmussen Centre of Excellence and the

scholarship to Mette Frandsen from the Danish Research council for independent research.

References

Andersen, T., Andersen, F. Ø. 2006. Effects of CO2 concentration on growth of filamentous algae and

Littorella uniflora in a Danish softwater lake. Aquatic Botany. 84(3): 267-271

Paper 2

18

Angers, D. A., Caron. 1998. Plant induced changes in soil structure: Processes and feedbacks. Development

in Biogeochemistry. 4: 55-72

Benoy, G. A., Kalff, J. 1999. Sediment accumulation and Pb burdens in submerged macrophyte beds.

Limnology and Oceanography. 44/4: 1081-1090

Blais, J. M., Kalff, J. 1995. The influence of lake morphometry on sediment focusing patterns. Limnology and

Oceanography. 40: 582-588

Bouwer, H. 2002. Artificial recharge of groundwater – hydrogeology and engineering. Hydrogeology Journal.

10: 121 – 142

Brunke, M. 1999. Colmation and depth filtration within streambeds: retention of particles in hyporheic

interstices. International Review of hydrobiology. Hydrobiology. 84: 99-117

Brunke, M., Gonser, T. 1997. The ecological significance of exchange processes between rivers and ground-

water. Freshwater Biology. 37: 1-33

Chambers, P. R., Kalff, J. 1987. Light and nutrient in the control of aquatic plant community structure I. In situ

experiments. The Journal of Ecology. 75: 611-619

Chen, X. 2000. Measurement of streambed hydraulic conductivity and its anisotropy. Environmental

Geology. 39(12): 1317-1324

Cherkauer D.S., Nader D.C. 1989. Distribution of groundwater seepage to large surface-water bodies - the

effect of hydraulic heterogeneities. Journal of Hydrology. 109: 151-165

Christensen, P. B., Sørensen, J. 1986. Temporal variation of denitrification activity in plant-covered littoral

sediment from Lake Hampen, Denmark. Applied environmental Microbiology. 1174-1179

Descloux, S., Datry, 0T., Philippe, M., Marmonier, P. 2010. Comparison of different techniques to asses

surface and subsurface streambed colmation with fine sediments. International review of

Hydrobiology. 95(6): 520-540

Fitts, C. R. 2013. Groundwater science. Oxford: Academic press Elsevier. 2nd.ed 63-68

Forster, S., Graf, G. 1995. Impact of irrigation on oxygen flux into the sediment: intermittent pumping by

Callianassa subterrranea and “piston-pumping” by Lanice conchilega. Marine biology. 123: 335-346


of aquatic macrophytes. Freshwater Biology. 57(5): 907-921

Förstner, W. C., Conradt, K. Jaksch, H., Schimkus, C., Schoer, J. 2008. Chemical speciation of heavy metals

in solid waste materials (sewage sludge, mining wastes, dredged materials, polluted sediments) by

sequential extraction. Universitätsbibliothek der Technischen Universität, Hamburg

Genereux, D., Bandopadhyay. 2001. Numerical investigation of lake seepage patterns: effects of porous

medium and lake bed properties. Journal of Hydrology. 241: 286-303

Gregg, W. W., Rose, F. L. 1985. Influence of aquatic macrophytes on invertebrate community structure, guild

structure, and microdistribution in streams. Hydrobiologia. 128: 45-56

Gyssels, G., Poesen, J., Bochet, E., Li, Y. 2005. Impact of plant roots on the resistance of soils to erosion by

water: a review. Progress in Physical Geography. 29 (2): 189-217

Paper 2

19

Hagerthey, S. E., Kerfoot, W. C. 1998. Groundwater flow influences the biomass and nutrient ratios of

epibenthic algae in a north temperate seepage lake. Limnology and Oceanography. 43(6): 1227-

1242

Hayashi, M., Rosenberry, D. O. 2005. Effects of groundwater exchange on the hydrology and ecology of

surface water. Groundwater. 40(3): 309-316

Hilton, J. 1985. A conceptual framework for predicting the occurrence of sediment focusing and sediment

redistribution in small lakes. Limnology and Oceanography. 30: 1131-1143

Huettel, M., Gust, G. 1992. Solute release mechanisms from confined sediment cores in stirred benthic

chambers and flume flows. Marine Ecology Progress series. 82: 187-197

Huettel, M., Ziebis, W., Forster, S. 1996. Flow-induced uptake of particulate matter in permeable sediments.

Limnology and Oceanography. 41(2): 309-322

Hutchinson, G. E. 1957. Concluding remarks. Cold Spring Habor Symposium. Quatitative Biology. 22: 415-

427

Iversen, J. 1936. Biologische Pflazentypen als hilfsmittel in der vegatationsforschung. Levin and Munksgaard

Iversen, T. M., Thorup, J., Hansen, T., Lodal, J., Olsen, J. 1985. Quantitative estimates and community

structure of invertebrates in macrophyte rich streams. Archiv fur Hydrobiologie. 102: 291-301

Karan, S., Kidmose, J., Engesgaard, P., Nilsson, B., Frandsen, M., Ommen, D. A .O., Flint, M. R., Andersen,

F. Ø., Pedersen, O. (2014). Role of a groundwater-Lake interface in controlling seepage of water

and nitrate. Journal of Hydrology. 517: 791-802

Kidmose, J., Engesgaard, P., Nilsson, B., Lair, T., Looms, M. C. 2011. Spatial distribution of seepage at a

flow-through lake: Lake Hampen, Western Denmark. Vadose Zone Journal. 10(1): 110-124

Kidmose, J,. Nilsson, B., Engesgaard, P., Frandsen, M., Karan, S., Landkildehus, F., Søndergaard, M.,

Jeppesen, E. 2013. Focused groundwater discharge of phosphorus to a eutrophic seepage lake

(Lake Væng, Denmark): implications for lake ecological state and restoration. Hydrogeology Journal.

21: 1787-1802. DOI 10.1007/s10040-013-1043-7

Kishel, H. F., Gerla, P.J. 2002. Characteristics of preferential flow and groundwater discharge to Shingobee

Lake, Minnesota, USA. Hydrological Processes. 16: 1921–1934

Landon, M. K., Rus, D. L., Harvey, F.E. 2001. Comparison of in-stream methods for measuring hydraulic

conductivity in sandy streambed attributes. Journal of Hydrology. 355: 164-180.

doi:10.1016/j.jhydrol.2008.03.018

Lee, D. R. 1977. A device for measuring seepage flux in lakes and estuaries. Limnology and Oceanography.

22: 140-147

Lehman, J. T. 1975. Reconstructing the rate of accumulation of lake sediment: The effect of sediment

focusing. Quaternary Research. 5: 541-550

Loeb, S. L., Hackley, S. H. 1988. The distribution of submerged macrophytes in Lake Tahoe, California and

Nevada, and the possible influence of groundwater seepage. Internationale Vereinigung fur

theoretische und angewante Limnologie, verhandlungen. 23: 1927 – 1933

Losee, W., Wetzel, R. G. 1993. Littoral flow rates within and around submerged macrophyte communities.

Freshwater Biology. 29: 7-17

Paper 2

20

Madsen, T. V., Cedergreen, N. 2002. Sources of nutrients to rooted submerged macrophytes growing in a

nutrient rich stream. Freshwater Biology. 47(2): 283-291

Mann, C. J., Wetzel, R. G. 2000. Hydrology of an impounded lotic wetland – subsurface hydrology.

Wetlands. 20(1): 33-47

Marmonier, P., Deletre, P. Y., Lefebvre, S., Guyon, J., Buolton, A. J. 2004. A simple technique using wooden

stakes to estimate vertical patterns of interstitial oxygenation in the bed of rivers. Archive Fur

Hydrobiologie. 160: 133-143


512

Meinikmann, K., Lewandowski, J., Nutzmann, G. 2013. Lacustrine groundwater discharge: Combined

determination of volumes and spatial patterns. Journal of Hydrology. 502; 202-211

Michaud, E., Desrosiers, G., Mermillod-Blondin, F. Sundby, B., Stora, G. 2005. The functional group

approach to bioturbation: The effects of biodiffusers and gallery-diffusers of the Macoma balthica

community on sediment oxygen uptake. Journal of Experimental Marine Biology and Ecology.

326(1): 77-88

Mitchell, A. R., Ellsworth, T. R., Meek, B. D. 1995. Effects of root systems on preferential flow in swelling soil.

Communication in Soil Science and Plant Analyses. 26: 2655-2666

Moeslund, B. 2000. Hampen sø – Environmental status 1971 – 1999. Vejle County (In Danish)

Møller, C. L., Kjøller, R., Sand-Jensen, K. 2013. Organic enrichment of sediments reduces arbuscular

mycorrhizal fungi in oligotrophic lake macrophytes. Freshwater Biology. 58: 769-779

Ommen, D.A., Kidmose, J., Karan S., Flindt, M. R., Engesgaard, P., Nilsson, B., Andersen, F.Ø. 2012.


Denmark. Ecohydrology, 5(3): 286-296

Palmer, M. A., Civich, A. P., Lake, S., Biro, P., Brooks, J. C., Dahm, C., Gibert, J., Goedkoop. W., Martens,

K., Verhoeven, J., Van De Bund, W. 2000. Linkage between aquatic sediment biota and life above

sediments as drivers of biological and ecological processes. Oxford journals. BioScience. 50(12):

1062-1075

Pedersen, O., Sand-Jensen, K., Revsbech, N. P. 1995. Diel pulses of O2 and CO2 in sandy lake sediments

inhabited by Lobelia dortmanna. Ecology. 76(5): 1536-1545

Petticrew, E. L., Kalff, J. 1991. Predictions of surficial sediment composition in the littoral zone of lakes.

Limnology and Oceanography. 36(2): 384-392

Raun, A. L., Borum, J., Sand-Jensen, K. 2010. Influence of sediment organic enrichment and water alkalinity

on growth of aquatic isoetid and elodeid macrophytes. Freshwater Biology. 55: 1891-1904

Rillig, M. C., Mummey, D. L. 2006. Mycorrhizas and soil structure. New Phytologist. 171: 41-53

Robe, W. E., Griffiths, H. 1998. Adaption for an amphibious life: changes in leaf morphology, growth rate,

carbon and nitrogen investment, and reproduction during adjustment to emersion by freshwater

macrophyte Littorella uniflora. New Phytologist. 140: 9-23

Roelofs, J. G.M., Schuurkes, J. A. A. R., Smits, A. J. M. 1984. Impact of acidification and eutrophication on

macrophyte communities in softwaters. II Experimental studies. Aquatic botany. 18: 389-411

Paper 2

21

Rosenberry, D.O., Toran, L., Nyquist, J.E. 2010. Effect of surficial disturbance on exchange between

groundwater and surface water in near shore margins, Water Recourses Research. 46(6): doi:

10.1029/2009WR008755

Sand-Jensen, K. 1998. Influence of submerged macrophytes on sediment composition and near-bed flow in

lowland streams. Freshwater Biology. 39: 663-679

Sand-Jensen, K., Borum, J., Binzer, T. 2005. Oxygen stress and reduced growth of Lobelia dortmanna in

sandy lake sediments subject to organic enrichment. Freshwater Biology. 50: 1034-1048

Sand-Jensen, K., Mebus, J. R. 1996. Fine-scale patterns of water velocity within marcrophyte patches in

streams. Oikos. 76(1): 169-180

Sand-Jensen, K., Prahl, C., Stokholm, H. 1982. Oxygen release from roots of submerged macrophytes.

Oikos. 38: 349-354

Schafran, G.C., Driscoll, C.T., 1993. Flow path-composition relationships for groundwater entering an acidic

lake. Water Resour. Res. 29, 145

Schuster, P.F., Reddy M.F., LaBaugh J.W., Parkhurst R. S., Rosenberry D.O., Winter T.C., Antweiler R.C,

Dean W.E. 2003. Characterization of surface water and groundwater movement in the littoral zone of

Williams Lake, a closed-basin lake in north central Minnesota. Hydrological Processes. 17: 823-838

Sebestyen, S. D., Schneider, R. L. 2004. Seepage patterns pore water and aquatic macrophytes:

hydrological and biogeochemical relationship in lakes. Biogeochemistry. 68: 383-409

Sebok, E., C. Duque, J. Kazmierczak, P. Engesgaard, B. Nilsson, S. Karan, and M. Frandsen (2013), High-

resolution distributed temperature sensing to detect seasonal groundwater discharge into Lake

Væng, Denmark, Water Resour. Res., 49, doi:10.1002/wrcr.20436

Sebok, E., C. Duque, P. Engesgaard, and E. Boegh, (2014). Spatial variability in streambed hydraulic

conductivity of contrasting stream morphologies: channel bend and straight channel, Hydrol.

Process. DOI: 10.1002/hyp.10170

Shaw, R. D., Prepas, E. E. 1990. Groundwater-lake interactions II. Nearshore seepage patterns and the

contribution of groundwater to lakes in central Alberta. Journal of Hydrology. 119: 121-136

Schälchli, U. 1992. The clogging of coarse gravel river beds by fine sediments, Hydrobiologia. 235(236):

189-197

Song, J., Chen, Xunhong., Cheng, C., Wang, D., Lackley, S., Xu, Z. 2009. Feasibility of grain-size analysis

methods for determination of vertical hydraulic conductivity of streambeds

Taylor, K., Wheatcraft, S., Hess, J., Hayworth, J., Molz, F. 1990. Evaluation of methods of determining the

vertical distribution of hydraulic conductivity. Groundwater. 37 (1): 88-98

Vestergaard, O., Sand-Jensen, K. 2000. Alkalinity and trophic state regulate aquatic plant distribution in

Danish lakes. Aquatic Botany. 67: 85-107

Vukovic M., Soro A. 1992. Determination of hydraulic conductivity of porous media from grain-size

composition, Water Resources Publication, Colorado, USA

Wetzel, R. G. 1990. Detritus, macrophytes and nutrient cycling in lakes. Scientific Perspectives in Theoretical

and Applied Limnology. 47: 233-249

Paper 2

22

Winkel, A., Borum, J. 2009. Use of CO2 by submerged rooted macrophytes. Annals of Botany. 103(7): 1015-

1023

Paper 3

Tacking groundwater flow during a flow reversal – nature’s

own tracer experiment


Paper 3

1

Tracking groundwater flow during a flow reversal – nature’s

own tracer experiment.

Mette Frandsen*, ^

, Peter Engesgaard+, Bertel Nilsson

^ Ole Pedersen

*

*The freshwater Biological Laboratory, Department of Biology, University of Copenhagen, Hillerød, Denmark


+Department of Geography and Geology, University of Copenhagen, Denmark

Abstract

Transient effects such as flow reversals are difficult to detect, but can have significant impact on lake

ecology. In this study we use a natural tracer, 18

O, to track the groundwater movement in the sediment of

Lake Hampen during and after a flow reversal, and on the same time estimating the effect on the nitrate

transport and denitrification.

Form our results we estimate that the flow reversal have been in effect from 2010 to 2012 – 2013.

The reason for the flow reversal was possibly a combined effect of reduced precipitation leading to lower

hydraulic heads in the catchment relative to lake stage. The precipitation decreased gradually from 913 mm

in 2007 to 704 mm in 2012. At the same time there was a drop in air temperature in 2009-20012 resulting in

precipitation as snow rather that rain leading to a lower infiltrating rate in the catchment.

During the flow reversal, lake water penetrated/seeped down through the sediment reaching a

depth of minimum 1.25 m below the lake bed. This was in an area normally discharging large amounts of

nitrate to the lake. During the flow reversal we found much lower than expected nitrate concentrations in the

sediment as the lake bed was saturated with lake water. As the flow reversal stopped and the directional flow

of groundwater reversed again, the nitrate concentrations increased again.

We examined the denitrification rates needed to reach the observed nitrate concentrations in 0.10

m depth. The denitrification rates required were 31 µmol N m-2

hour-1

which is in the lower end of what other

studies found for the same area.

When the experimental period was over, the lake had not fully recovered as indicated by δ18

O and

nitrate concentrations in the sediment. We speculate that the system possible have undergone several flow

reversals in historical time, as the causes for this flow reversal were relative small.

Paper 3

2

Introduction

Lakes are integral parts of larger groundwater flow systems and must be considered open systems reacting

to and affecting the groundwater through exchange of water over the sediment water interface (SWI) (Sear et

al., 1999; Hancock et al., 2005). The discharging groundwater supplies the lake with often high

concentrations of solutes and inorganic carbon accumulated in the catchments. This can benefit the biota in

the discharge zones (Frandsen et al., 2012; Hagerthy & Kerfoot, 1989), but can also have a substantial

impact on the trophic status of the lake (Sebestyen & Schneider, 2004; Hayashi & Rosenberry, 2002;

Hagerthey & Kerfoot, 1998). In some cases, groundwater input of nutrients accounts to up to 50% of the

annual nutrient load to a lake (Brock et al., 1982; Shaw & Prepas, 1989). This makes the understanding of

groundwater flow in lake settings an important component in order to understand the ecology of the system.

The groundwater systems are overall controlled by recharge from precipitation, evapotranspiration,

and exchange of water with lakes. On this scale the directional subsurface flow of groundwater is mainly

controlled by topography (gravity) and geology. The groundwater will flow from high elevations to low

elevations, driven by differences in potential energy. On a smaller scale the flow pattern is affected by

sedimentary heterogeneity, causing differences in hydraulic conductivity, hereby affecting the groundwater

flow.

In perfectly homogeneous, isotropic and steady state systems the groundwater discharging to a

lake will often decrease exponentially with the distance from the lake shore (McBride & Pfannkuch, 1975;

Pfannkuch & Winter, 1984; Lee, 1980). Deviations from this flow pattern will typically be due to the aquifer-

lake geometry (Winter, 1999), anisotropy in the sediments (Pfannkuch & Winter, 1984), geological

heterogeneity (Tóth, 1999) and seasonal changes in precipitation and drainage patterns (Downing &

Peterka, 1978; Rosenberry & Winter, 1997). The sensitivity of the groundwater surface water flow systems to

these factors, gives rise to a number of transient effects (Cheng & Anderson, 1994). Of the more important

effects are flow reversals. During a flow reversal, the directional flow of the groundwater will change either

locally or in the whole groundwater – lake system. In the extreme case this can change a discharge site, to a

recharge site (Ala-aho et al., 2013).

Flow reversals are typically caused by the creation of groundwater mounds or simply as a result of

changes in the regional and local in- and output of water to the system (LaBaugh et al., 1987; Rosenberry &

Winter, 1997). Even though many studies neglect transient effects (Winter, 1978; Cherkauer & Hensel, 1986;

Krabbenhoft et al., 1990) many studies point to the importance of these effects in altering the groundwater

flow pattern (Anderson & Cheng, 1994; LaBaugh et al., 1987).

Meyboom (1967) was one of the first to describe the importance of transient effects on the

groundwater flow patterns around lakes by showing that the groundwater flow was seasonally determined

with long periods of directionally diverted flow. Later it was described how the formations of groundwater

mounds could create stagnation points for groundwater (Anderson & Munter, 1981; Winter, 1983) and how

the creation of groundwater mounds could be seasonal (Cherkauer & Zager, 1989; Winter, 1986). Downing &

Peterka (1978) showed that groundwater discharge increased with increased rainfall, Rosenberry & Winter

Paper 3

3

(1997) showed how an increase in precipitation could be measured directly as rapid increase in heads, and

Sacks et al. (1992) used a modeling approach to show how high rainfall could cause local flow reversals

changing both recharge areas into discharge areas and vice versa.

Flow reversal will naturally also impact the solute in- and output to a lake (Anderson & Cheng,

1993; Steinwand & Richardson, 1989) and could under some circumstances have a great impact on the lake

ecology even though it is unknown if this can result in permanent changes in the lake trophic status over

time.

As the groundwater normally discharges in the littoral zone, the biota there benefits from the

groundwater supply of inorganic carbon and nutrients. During a flow reversal in these settings, lake water will

instead seep downwards through the sediment carrying little nutrient and less inorganic carbon than ground

water normally does. This will affect the sediment associated species as shown by Frandsen et al. (2012),

who shows that the growth rates of small isoetid plant species are significantly affected by the groundwater

seepage, and by Hagerthy and Kerfoot (1998) who shows the same effects for algae. It could be speculated

that in highly competitive communities, a longer period of flow reversal could affect the species composition

in the affected zones. This however is unknown.

Flow reversal can be difficult to detect as they are normally transient and sometimes very local.

However, such events are interesting, not only in an ecological sense but also in a hydrological sense. By

using natural tracers to follow the changes in directional flow during flow reversals, they can utilize a better

understand groundwater-lake interactions. The objectives of this study was therefore; (1) to demonstrate how

natural stable oxygen isotopes and reactive nitrate tracers can be used to track a flow reversal at Lake

Hampen, Denmark during more than two years and (2) to estimate denitrification rates at the groundwater-

lake interface during flow reversal..

Materials and methods

Study site

Lake Hampen is located in the western part of Denmark in the middle of the Jutland peninsula just east of

the last glacial advance (figure 1). The lake has a surface area of 0.76 km2 and is situated in a 15-25 meter

deep layer of coarse melt water sands and gravel (Kidmose et al., 2011).

The lake is characterized as a flow through lake, and under normal conditions the directional flow

of the groundwater goes from north-east, east and south-east to west direction. Approximately 2/3 of the

water is received from groundwater discharge and 3/4 of the water leaves the lake through groundwater

recharge at the western shore line. There are no major in- or outlets so precipitation and evaporation account

for the remaining in- and out-fluxes. Using a Darcy approach Ommen et al. (2012) estimated a total

discharge of 4500 m3 day

-1 with fluxes ranging between 44 – 134 L m

-2 day

-1 in the area investigated in this

study.

Paper 3

4

The catchment (993 ha) is primarily covered with forest (62%) and agricultural land (30%).

Experiments were placed near the agricultural land located on the North-Eastern shoreline where most of the

groundwater discharges the aquifer (Ommen et al., 2012).

Instrumentation

In November 2010 a 5*5 matrix covering an area of 20*20 meter of small pore water seepers (N=30) were

installed in the lake bed on the North-Eastern shore. The seepers were installed 0.10, 0.25, 0.50. 0.75, 1.00,

1.25 meters below the sediment. Matrix rows were placed in lines so each line perpendicular and parallel to

the shore line contained a pixi-seeper in each of these depths, thus covering a larger area, but also enabling

a projection of the results into a 1D transect line (figure 1 and 2). The filters were installed by hammering

them down using a fitted steel pipe and a hammer. The filters were connected to polyvinylchloride (PVC)

tubes enabling sampling of pore water from land. The PVC tubes were secured along the lake bed and

mounted on land. Each PVC tube was connected to a three-way valve for sampling.

In August 2011 an additional line of deeper wells (A1-A6) were installed perpendicular to the shore

line with four on-land and two off-shore. The wells were installed in the following depths; A1 1.1 m, A2 1.1 m,

A3 3.6m, A4 3.2 m, A5 5.2m A6 7.2 m, and A7 5.5m.

Figure 1. Map of Lake Hampen showing the instrumentation of the North-Eastern shoreline where groundwater under normal conditions discharge to the lake. The map shows location of the seepers, the A-Wells and the W-Wells. On the horizontal axis, negative values are in the lake and positive values are on land. The stippled line shows the shoreline and the dotted line indicates the 1D projection shown in Figure 2.

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5

In February 2013 three deeper wells (W-wells) were installed in 16 m (W1), 14 m (W2) and 6.4 m

(W3) depths (figure 2). Wells W2 and W3 were located 7 and 22 meter off-shore, respectively. These wells

were constructed of galvanized steel pipes with an inner diameter of 2.1 (A wells) or 2.7 cm (W-wells) and a

screen length of 9 cm. The wells were installed using a pneumatic hammer.

Oxygen isotopes (δ 18

O)

The flow reversal was followed using δ 18

O. Oxygen isotopes can be used to track groundwater - surface

water interactions in systems exhibiting a distinct contrast in the δ 18

O concentrations between groundwater

and surface water (Krabbenhoft et al., 1990). The groundwater will attain the average concentration of the

precipitation and this value is often much lower than the concentration in the surface water due the

evaporative fractionation processes in the surface water. The difference in the concentration of the heavy

oxygen isotope will be a proxy for the origin of the water (i.e. groundwater or surface water).

The fractionation is calculated by comparing the sample with the Vienna standard mean ocean

water value (VSMOWV) (Appelo & Postma, 2005) as described in equation (1)

( ) ( )

( ) (1)

2007 2008 2009 2010 2011 2012 2013

January 155 106 42 14 51 92 69 February 82 63 27 26 41 33 26 March 60 105 49 32 22 34 8 April 12 53 9 21 22 71 29 May 60 7 59 61 45 40 82 June 100 42 41 50 84 104 61 July 119 56 100 93 90 111 16 August 50 151 76 137 127 42 49 September 99 66 59 79 118 124 92 October 40 127 98 78 65 118 129 November 56 77 147 90 23 70 63 December 81 22 60 23 111 76 133

Total 913 875 757 704 798 915 756

Table1. Monthly total precipitation (mm) from 2007 - 2013. Data are from The Danish Meteorological Institute (DMI).

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6

The difference in δ18

O concentration between groundwater and surface water depends strongly on the

residence time of water in the lake. Lakes with long residence times will have a higher fractionation than

lakes with short residence times. Lake Hampen, exhibits a distinct difference in the δ 18

O concentration in the

groundwater and the surface water. Kidmose et al. (2011) found that the δ18

O in Lake Hampen is in the

range of -4.5 to -3.3 ‰ and approximately -8.2 ‰ for groundwater.

During a two year period from 2012 to 2013 water samples from the seepers were collected

monthly for δ18

O analysis. The samples were collected by connecting a syringe to the valves attached to the

PVC tubes connected to the seepers. Water samples were collected after clean pumping (≥ 3 times the

volume of the pipe + seeper). Surface water samples for δ 18

O analysis were collected directly ~ 8 meters

from shore. The samples were kept in air free bottles < 5° C until analysis.

The samples were analysed with a Piccaro Water analyzer. Samples were analysed 6 times and a

standard deviation were calculated for each sample.

To more clearly identify the movement of water beneath the lake bed during and after the flow

reversal we used the δ 18

O concentration, where half the water was from groundwater and surface water

respectively. This value averaged to δ 18

O = -5.85 ‰.

Nitrate

As part of the monitoring we collected water samples in the seepers and the A and W-weels for nitrate

analysis during 2011 – 2014. Samples from seepers were collected as described above. Samples from wells

Table 2. Advective and diffusive nitrate flux assuming nitrate concentration of 60 mg L-1

in 1.25 m depth and 0.5 mg L

-1 in 0.10 m depth.

Table 3. Nitrate concentration in the A-wells from October 2012 to June 2014.

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7

were collected after clean pumping three times using a peristaltic pump. All samples were kept in bottles < 5°

C until analysis. Samples not immediately analysed was freezed within 24 hours. All samples were filtrated

before analysis. Dissolved nitrate was analysed spectrophotometrically on an automated ion flow injection

analyzer (QuickChem methods 10-107-04-1-C).

To estimate the expected nitrate flux in the top sediment a combined advection-diffusion transport

equation was used (equation 3)

F = -nD * dC/dx + qC (3)

Where (mg m-2

day-1

) is the mass flux, n (%) is the porosity, D is the molecular diffusion coefficient,

dC/dx is the concentration gradient over a sediment thickness L, q is the seepage flux in m day-1

. Here we

used molecular diffusion coefficient from Rittmann and Manem (1991) of 1.4*10-4

(m2 day

-1), a porosity of

0.35, and a sediment thickness of L=115 cm (from 10 cm to 125 cm depth).

Seepage meters

A line of 7 seepage meters were deployed in the lake bed (figure 1). Seepage meters were constructed from

the bottom of a steel drum (internal diameter, 57 cm) slightly modified from the design of Lee et al. (1977).

To avoid frictional errors (Rosenberry, 2005) a pipe with a large diameter (i.d. 1.5 cm) was connected to the

seepage meter as suggested by Fellows and Brezonick (1980). A Stotz fitting was connected to the pipe

allowing easy connection to a valve, also with a Stotz fitting, to which a 4 L bag was attached. A rigid plexi-

glas box covered the seepage meter to protect the bag from currents and water movement disturbing the

measurements (Libelo & MacIntyre, 1994). Before any measurements were carried out, the seepage meter

was left in the lake bed to equilibrate to obtain the same hydraulic pressure inside the seepage meter as

outside and the sediments to settle around the seepage meter. Before an experiment the bag was pre-filled

with 1 L of lake water to avoid initial short-term fluxes (Shaw & Prepas, 1989), excessive air was forced out,

and the bag weighed. When the bag was connected to the pipe, the valve was opened, the time recorded,

and the rigid box placed on top of the seepage meter. At the end of the experiment, time was recorded and

the bag weighed again. The discharge, q, was calculated as q=V/(t*A), where V is the change in volume

over time period t, and A is the area of the seepage meter (~ 0.25 m2).

The seepage rates was measured two times in April 2014 and should only be seen as a snap shot

of the flow condition near the end of this experiment.

Climatic data

Data on precipitation and temperature is obtained from Danish Meteorological Institute. The data represent

monthly averages from 2007 – 2013.

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8

Results

Figure 4 shows the lake stage development from 2007 to 2013 (no data in 2009). In 2007-2008 the lake

stage was fairly stable followed by a sudden drop in 2010 to 2011. During 2012 the lake stage increases

again reaching close to normal level in 2013 (figure 4).

Figure 3 shows the head distribution during the flow reversal and under normal flow conditions.

During the flow reversal the head was similar to or sometimes lower than that of the lake, giving almost

stagnant flow or recharge of water from the lake. The head distribution in the near shore areas was

congruent with the lake stage with very low water levels in the lake during the flow reversal occurring in

2010-2011, where the lake stage was unusually low.

The precipitation followed a similar pattern. In 2007-2008 during normal flow conditions, the

precipitation was around 900 mm. During 2010 and 2011, the precipitation drops to 700 mm and 800 mm

respectively. In 2012 as the precipitation increases again and the lake stage increases (figure 4, table 1).

The average air temperature drops slowly during the 2007-2010 reaching a minimum in 2010

where the lowest minimum temperature of minus ~-23 C° was recorded. After 2010 the average temperature

increases slowly again until 2013 (Figure 5). In many of the months between 2012 and 2011 the average

winter temperature was 0 C°, why the lake was ice-covered and the precipitation fell as snow.

Oxygen isotopes

in the top 25 cm of the sediment (seepers in 0.10 and 0.25 m depth) the δ 18

O signal reflect lake water (> -

5.85 ‰) over the entire period from 2012 to 2013 (figure 6). In 0.50, 0.75, 1.00 and 1.25 m depth, lake water

Figure 2. 1D projection of the transect showing depth distribution of A-wells, W-wells and seepers. The depth is shown as meters above sea-level. On the horizontal axis, negative values are in the lake and positive values are on land

Paper 3

9

was found primarily in the beginning of the period, with a groundwater signal increasing from October 2012

to October 2013 (figure 6). In 1.00 and 1.25 m depth almost only groundwater is found in September 2013.

This is also obvious from figure 7. In period 1 (March to June 2012) there is almost no variability in

δ 18

O from the top of the sediment to the deepest placed seepers. The signal was steady around -3.5 ‰

reflecting lake water. This is also the value we have used to calculate the half-lake-water and half-

groundwater composition as it better represents a long-term average value for lake water. In period 2 (July –

November 2012) there is little variability around -2.9 ‰ in the top 0.25 m of the sediment which is

approximately equal to lake water (-3.5 ‰). Below 0.25 m some variability is observed and the δ 18

O signal

increases from 0.25 m to 1.25 m where the signal reached an average of -5.1 ‰ which is almost the

composition of equal parts of lake water and ground water. In period 3 (September 2013) the average δ 18

O

was close to lake water in the top 0.25 m of the sediment followed by a distinct increase deeper in the

sediment and a signal almost equal to groundwater in the deepest placed seepers.

Figure 8 shows the time series of the δ 18

O signal in all seepers in five zones (see figure 9) with

increasing distance to shore. In zone 1 nearest to the shore line the top 0.25 m had a δ 18

O signal similar to

lake water over the entire period. Below 0.25 m the signal reached that of groundwater or close to

groundwater in May-July 2012. In zone 2 all seepers except in 1.25 and in 0.75 m show lake water signal

until December 2012. The δ 18

O signal decreased over the entire period in 0.75 m depth reaching close to

groundwater signal in December 2012. The seeper in 1.25 m had a stable lake water signal until November

2012, then dropping quickly to groundwater signal in December 2012. The seeper in 1.00 m reflected lake

Figure 3) Head distribution (m) measured in A-wells (se figure 2) in November 2011 indicating low to no groundwater discharge to the lake (values in brackets) and again in both A-well and W-wells in April 2013 indicating close to normal groundwater discharge to the lake. Stippled lines show isopotential map based on head distribution in 2013.

Paper 3

10

water until the last measurement in September 2013 where the signal returned to groundwater signal. In

zone 3 and 4 all seepers had stable lake water signals until December 2012. In zone 3 this was followed by

a decrease in the δ 18

O signal in the three deepest seepers (0.75, 1.00 and 1.25 m depth) reaching close to

groundwater signal in September 2013. In zone 4 the stable period was followed by a decrease in the four

deepest seepers reaching groundwater signal in the two deepest seepers and close to groundwater signal in

0.75 and 0.50 m. In zone 5 the δ 18

O signal was also stable during 2012, however some variability was

observed in the 1.25 m depth, where the signal decreased to -5.3 ‰ in July 2012 followed by a return to lake

water signal during the following month. In September 2013 only the two deepest seepers reflected

groundwater signal.

Figures 10, a-d show a spatial representation of the δ 18

O distribution and the groundwater-surface

water-front (GS-front, solid heavy line), where the δ 18

O signal is half groundwater and half surface water (= -

5.85 ‰). In figure 9, a (April 2012), the GS-front was only visible in the near shore area, zone 1, and was

located below 1.00 m in the sediment. In zone 2-5 the δ 18

O reflected lake water or close to lake water in all

depths with a slightly decreasing signal nearer the shore in zone 1.

Four months later in August 2012 the GS-front had moved upward and outwards and was now

apparent in 0.50 m below the sediment in zone 1 the near shore area. This was also reflected in the

distribution with a slightly decreasing signal in zone 1 and 2 (Figure 9, b). Three months later in November

2012 the front had moved even further off-shore and was visible in both the near shore area at 0.50 m depth

Figure 4. Lake stage and precipitation from 2007 - 2013. Lake stage is shown as meters above sea level (m). Horizontal stippled lines indicate mean lake stage for the given year. Black horizontal bars indicate mean annual precipitation. Vertical stippled line indicates absence of data from 2009. Precipitation data are from The Danish Meteorological Institute (DMI).

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11

and also in 1.00 – 1.25 m depth in zone 2. The δ 18

O signal had decreased in all zones in the top seepers

indicating an outward and upwards movements of the groundwater (Figure 9, c). Finally in September 2013,

13 months later, the front was apparent in all zones up to ~ 0.50 m below the sediment in zone 1 and to 1 m

below the sediment in zone 5 (figure 9, d). The δ 18

O in the shallow seepers was ~-4 ‰ compared to the

initial value in April 2012 of ~ -2.6 ‰

The overall movement of the GS-fronts can be seen in figure 10. The GS fronts clearly move

upwards and outwards over the experimental period as the effect of the groundwater reversal diminishes and

the groundwater returns to the sediment below the lake again.

Figure 5. Monthly minimum, maximum and mean winter air temperature (monthly average) from 2007 - 2013.

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12

Nitrate

Normally the lake receives large

amounts of nitrate with groundwater

discharge (Kidmose et al. 2011).

However, during the flow reversal

lake water penetrated the sediment

reaching a depth of at least 1.25 m.

(figure 9, 11). Figure 11 shows the

average nitrate concentration in the

different depths over the

investigation period. In the top 0.10

m of the sediment the variability in

concentration is large, whereas it is

more stable deeper in the sediment.

In December 2012 the nitrate

concentrations suddenly increase in

seepers embedded deepest in the

sediment as it was also observed

with the δ 18

O signal.

Separating the data into four periods based on the development seen in figure 11 the trends

become more obvious (figure 12). In the first two periods the variability with depth is very small with

variations around a few mg L-1

(notice that the periods are not the same as for δ 18

O). In the third period

(June-December, 2012) a steep

increase below 0.50 m reaching

concentration of ~7-8 mg L-1

is

seen. In the last period, only

representing June 2014, the

increase below 0.50 m is still

obvious. In 0.75 m the

concentrations have increased

further, but in the two deepest

seepers the concentrations have

decreased compared to the

previous period. Some of the

variability is lost in this period,

as the data points are averages

from all the seepers in a given

Figure 6. δ 18

O signal at the different depths from March 2012 to September 2013. Each data point is the mean δ

18O value

of all seepers at a given depth at a given time. The two horizontal lines indicate the δ

18O signal for lake water (~ -3.5

‰) and groundwater (~ -8.2 ‰). The stippled line indicate the δ

18O signal of half lake water half groundwater (~-5.85 ‰).

Figure 7. δ 18

O (‰) with SD (n = 4 -20) in seepers under the transect. Each data point represents the mean δ

18O

signal value from all seepers at a given depth at a given period.

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13

depth regardless of distance to shore.

In figure 13 the development in the 5

zones shows more details. In zone 1 a steep

increase in nitrate concentration is seen from April

2012 in all seepers placed deeper than 0.25 m in

the sediment. The concentration increase levels out

in the seepers around August 2012 having reached

~60 mg NO3 L-1

in the two deepest seepers (1.00

and 1.25 m). In June 2014 the nitrate

concentrations suddenly drops again to values

close to zero. In zone 2 some of the same

dynamics is seen. Until August 2012 the

concentration are very low in all depths. After

August 2012 the nitrate concentrations in the

deepest seepers start to increase reaching 12.9 mg

NO3 L-1

in December 2013. In June 2014 the same

drop in concentration is seen as in zone 1.

In zone 3 there is some variability in 0.75

m but in general the concentrations are very low in

all depths over the whole period from 2011-2014.

The concentrations in zone 4 a and 5 are much

more variable and with generally higher

concentrations. In zone 4 there is more variability

especially in 0.10, 0.25. 0.50 and 0.75 m below the

sediment and two concentration peaks are

observed around October 2011 and again in June

2012, but only in 0.25 m. In the top layer of the

sediment (0.10 m) the variability is high and the

concentration varies between 0 and 10 mg NO3 L-1

.

In June 2014 the concentrations in the two deepest

seepers, 0.75 and 1.25m, increase reaching 21.2

and 12.5 mg NO3 L-1

respectively. In zone 5 the

variability is large over the whole period, but the

concentrations do not exceed 12 mg L-1

with the

exception of the seeper placed in the top 0.10 m,

where a peak reaching 19.5 mg NO3 L-1

is seen in

June 2012.

Figure 8. The δ 18O development with time in

zones 1-5 of various distances to the shore line

(see figure 8). Zone 1 is the transect line

nearest to the shore and zone 5 is the transect

line furthest from shore. Each data point shows

the δ 18O signal from a single seeper during

the periode.

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14

The Nitrate flux in the sediment was calculated using the combined advection-diffusion equation

Figure 10. Cross section of transect with projection of all seepers. Distance between each

row of seepers was 5 m and each row was sampled for every 0.25 m depth. The contour

lines show the δ 18O signal and the bold black line indicates the GS-front (front of water

where the δ 18O signal is 50% groundwater and 50% lake water = - 5.85 ‰). Measured

in 24-04-2012 (a), 18-08-2012 (b), 12-11-2012 (c) and 20-09-2013 (d).

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15

using average measured fluxes (q) from the two seepage meter measurements (0.3 cm/day). We found

nitrate concentration ~60 mg L-1

in 1.25 m depth (zone 1, last period, Figure 13) and low concentration

nitrate in 0.10 m ~0.5 mg L-1. These values were used to calculate the diffusive and the advective nitrate

fluxes. The calculated advective and diffusive fluxes were 201 and 2.5 mg NO3 m-2

L-1

d-1

respectively (table

2).

Discussion

During this study we have captured the back seeping of groundwater after a flow reversal in Lake Hampen

using primarily data on δ 18

O. Prior to our study, lake water penetrated the lake sediment down to a depth of

more than 1.25 m and during 2011 – 2013 the groundwater gradually returned.

Several studies have examined the north-eastern part of the lake since 2007. General discharge

rates near the transect have previously been reported to be between 0.1 – 6.9 cm day-1

(Kidmose et al.

Figure 10. Cross section of transect with projection of all seepers. There is 5 meter between each

zone and 25 cm between each seeper in depth. The grey lines show the front where the δ 18O

signal is 50% groundwater and 50% lake water (δ 18O = -5.85 0%) from March 2012 to

September 2012. The line numbers corresponds to the following dates: 1) 22 Mar 2012, 2) 24 Apr

2012, 3) 16 Jun 2012, 4) 14 Aug 2012, 5) 24 Sep 2012, 6) 10 Oct 2012; 7) 12 Nov 2012 and 8)

20 Sep 2013.

Paper 3

16

2011; Ommen et al. 2012; Karan et al.

2014). Thus, a gradient towards the lake

is normally observed. In this study we

found a lower than expected

groundwater head gradient in the near

shore area of the transect (figure 3). In

2011 the head was low or similar to the

lake stage (figure 3) and under these

conditions the water slowly recharges

the aquifer, but more importantly, at

these low fluxes, lake water diffuses into

the lake bed. In 2013 the head gradient

was again closer to normal indicating

groundwater discharge again (figure 3).

In April 2013 we also measure seepage

rates within the normal range, although

in the low end (0.3 cm/day). This

suggests that the flow reversal have

been in effect from at least 2010 to around 2012-2013.

The lake stage was followed from 2007 – 2013. In 2010 and 2011, where we expected the flow

reversal to be in effect, the lake stage was continuously under the average lake stage for the whole period

(figure 4). The period of low water table is congruent with the low head measured in the wells and the low

lake stage seems to be coupled to the flow reversal.

From previous studies we know that the δ 18

O signal in the Lake Hampen is approximately -3.5 ‰

whereas the δ18

O signal of the groundwater in that area is approximately -8.5‰ (Kidmose et al. 2011).

Given this information we estimated that during the flow reversal the lake water had penetrated the lake

sediment down to a depth of at least 1.25 m (the deepest sepeers) (Figures 7,8,9).

Figure 8 shows the overall development in δ18

O in three distinct periods. In period 1 (March – June

2012) we registered lake water in all the seepers. There was little variation between the deepest seepers and

the seepers installed just below the sediment surface. This was under the full effect of the flow reversal and

all the water in the lake sediment was lake water. The first sign of a return to normal flow conditions was

visible in period 2 (July – November 2012), where the variability and the δ18

O concentration increased as the

groundwater returned. In period 3 (September 2013) we see a marked increase in the δ18

O concentration

below 0.25 m and we see a clear groundwater signal in the deepest seepers (figure 7, period 3).

However, the returning groundwater mainly occurred in the near-shore zone (figure 8) where the

groundwater was visible in all the seepers already in august 2012. Further from shore the groundwater signal

was delayed compared to zone 1 and the signal was mainly seen in the deepest seepers. This return pattern

was expected as the normal flow pattern of groundwater discharge to a lake will decrease with distance to

shore and because of normal heterogeneity in the sediment.

Figure 11. Nitrate concentrations at the different depth from

March 2012 to June 2014. Each data point is the mean

nitrate concentration of all seepers in a given depth at a given

time.

Paper 3

17

By assuming uniform flow of the groundwater

perpendicular to the lake shore, we plotted the movement of the

GS-front, represented by the δ 18

O concentration that was half

groundwater half lake water (Figure 9). From the δ 18

O data it is

evident that the system is returning to normal condition after a

flow reversal. The re-emerging groundwater infiltrated the near

shore seepers first and emerged upwards and outwards from

there. At the end of the investigation period, the GS-front water

had re-emerged up through the sediment to around 0.75 m

below the sediment surface (Figure 9). Groundwater was visible

up to 0.50 m below the sediment surface in the near shore zone

(zone 1), but only up to 1.00 m in the other zones further from

shore (figures 8, 9, 11)

The conditions are not back to what we would expect

during normal flow conditions. Karan et al. (2014) measured the

δ 18

O signal in the top 0.50 m of the sediment at a neighboring

site using a diffusion sampler (data from 2007). They found δ

18O concentration of -7.5 ‰ in 0.10 m with a decrease up to -

8.2‰ in 0.40 m below the sediment. In contrast we find lake

water signal in the top 0.25 – 0.50 m of the sediment and only

groundwater in ~ 0.50 m in the near shore zone. Either the

return to normal flow conditions is not finished or the seepage

rates are just lower here (as indicated by the seepage meter

measurements).

The flow reversal was also evident from the nitrate

data. Despite the more complex behavior of nitrate caused by

seasonally determined biological activity in the catchment and

the lake bed sediment, some interesting patterns are still

obvious. In figure 11 we see a clear deviation from the expected nitrate concentration with extremely low

concentrations all the way down to 1.25 m below the sediment in almost all the seepers. This is obviously

lake water that has seeped/diffused down through the sediment during the full effect of the flow reversal. The

low concentrations are evident both during summer and winter months, eliminating the seasonally variability

as a viable explanation. After almost two years of monitoring low concentrations, we suddenly see a

significant increase during 2012. Still, the concentrations are much lower compared to what other studies

have found (see below) (figure 12).

The returning Nitrate with the returning groundwater indicates a slow recovery to normal conditions

especially in the zone closest to the shore line (figure 13). However, the sampling in June 2014 shows low

concentration in zone 1. In the deeper A-wells the concentration have increased steadily since 2012 (Table

3) and in these deep wells the concentrations are also low in 2014 explaining the suddenly low concentration

Figure 12. Nitrate concentration with

SD (n = 4 - 20) in seepers under the

transect. Each data point represents

the mean nitrate concentration from

all seepers at a given depth at a given

period.

Paper 3

18

in the seepers at the same time. There could be some

seasonal effects in play, but the big drop in

concentration is more likely caused by a pocket of

nitrate poor water moving towards the lake from the

catchment. The zone closet to shore react more

quickly than the zone further off shore, explaining the

different effect on the signal between zone 1 and the

other zones.

Ommen et al. (2012) estimates a total input

of nitrate-N around 3000 kg year-1

with the majority

coming from the agricultural transect. They measured

the nitrate concentration in the pore water in the

sediment and found average concentrations 86.8 mg L-

1 in 0.10 – 0.25 m; 31 mg L

-1 in 0.50 m, 86.2 mg L

-1 in

1.50 m. This is significantly higher than the

concentrations found in this study and this strongly

indicates that the conditions present before the flow

reversal is not reached yet. Still in 2012, we saw nitrate

concentration up to ~60 mg L-1

in the deepest seepers

(Figure 13, zone 1). However, in 0.10 m our highest

measured value is 20 mg L-1

and the average in each

zone is much lower. In the last sampling the overall

average nitrate concentration in 0.10 m was only ~8

mg L-1

.

Given the discharge rate we calculated using

seepage meter measurements and the nitrate

concentration measured in the deep seepers we

calculated a nitrate flux to 203 mg m-2

d-1

assuming

nitrate concentration of 60 mg L-1

in 1.25 m depth and

nitrate concentration of 0.5 mg L-1

in 0.10 m depth.

Biotic uptake and denitrification in the oxygenated part

of the sediment is responsible for reducing the nitrate

concentrations from 60 mg L-1

to 0.5 mg L-1

(Christensen & Sørensen, 1986; Hill et al., 2000;

Pfenning & McMahon, 1996; Ommen et al., 2012).

Assuming that denitrification removes the all excess

nitrate before it reaches the top of the sediment

denitrification rates of 31 µmol N hour-1

m-2

would be

Figure 13. The nitrate concentrations over time

in zones 1-5 of various distances to the shore

line (see figure 8). Zone 1 is the transect line

nearest to the shore and zone 5 is the transect

line furthest from shore. Each data point

shows the nitrate concentration from a single

seeper during the period.

Paper 3

19

needed. In comparison Christensen and Sørensen (1986) found denitrification rates in Lake Hampen ~50

µmol N hour-1

m-2

for a system with low nitrate concentrations. However, when supplying nitrate, they saw a

7-fold increase in denitrification activity reaching rates of 225 – 350 µmol N h-1

m-2

. Given this, our

denitrification rates are low and they will possibly increase as the nitrate returns to the system.

In this study we have a long term flow reversal probably caused by meteorological factors.

First of all we observe a falling lake stage and groundwater head in the near shore areas in 2010

and 2011. The total precipitation decreased from 913 mm in 2007 and 875 mm in 2008 to 754 and 704 mm

in 2009 and 2010, respectively. This could affect both the lake stage directly but also the groundwater head,

due to lower infiltration (Downing and Peterka, 1978; Sacks et al., 1992; Rosenberry & Winter, 1997). At the

same time, the air temperature in the winter period dropped in 2009 and 2010 why a larger proportion of the

precipitation in those years fell as snow, with a following lower infiltration rate.

However, none of the meteorological data are very extreme seen in a historical time frame, and if

these factors lead to a groundwater reversal in Lake Hampen, it should be anticipated that the lake possibly

have undergone several groundwater reversals back in time.

When we refer to the normal flow conditions in the lake by inspecting the results from the more

extensive monitoring of the lake that started in 2007 with Kidmose et al. (2011), Ommen et al. (2012) and

Karan et al. (2014), we might be mistaken. During this period the climate was relatively wet and warm. The

head gradient in this period was higher. Hence, the findings from this period could give a skewed idea of the

normal flow conditions at the lake. Given all these findings we must conclude that the hydrology in and

around Lake Hampen are more complex than we previously expected. Temporal and spatial variation in the

in- and outputs are quite significant and to fully understand the system, data on a historical time scale as well

as on different spatial scales are needed.

A flow reversal can have an impact on the ecology of the system. In freshwater lakes the chemistry

in the surface water can be seen as an integrated response to all the catchment specific in- and outputs as

well as the in-lake processes (Krabbenhoft & Webster 1995; Gurrieri & Furniss 2004; Rimmer et al., 2006).

Groundwater seepage discharge can have a significant effect on the plant community associated with the

local discharge zones as shown by Frandsen et al. (2012). During a flow reversal plants benefiting from the

groundwater supply of nutrients will be affected. On a larger scale prolonged flow reversal might affect the

ion-balance of the entire system, especially in systems where the in- and outputs mainly takes place through

groundwater seepage hereby affecting not only the littoral communities associated with the sedimentary

conditions but also the planktonic species in the surface water.

Acknowledgements

This project was founded by The Danish Council for Independent Research – Nature and Universe. We like

to thank Mithra Christin Hajati and Kristian Färkkilä Knudsen and Carlos Duque Calvache for valuable help.

Paper 3

20

References

Ala-Aho, P., P.M. Rossi, and B. Kløve, 2013, Interaction of esker groundwater with headwater lakes and

streams, Journal of hydrology. 500: 144-156

Appelo, C.A. J. & Postma, D. Geochemistry, Groundwater and pollution. 2005. A.A. Balkema publishers.

458-462

Anderson, M. P., Cheng, X. 1993. Long. And short-term transience in a groundwater/lake system in

Wisconsin, USA. Journal of Hydrology. 145: 1-18

Anderson, M. P., Munter, J. A. 1981. Seasonal reversals of groundwater flow around lakes and the relevance

to stagnation points and lake budgets. Water. Resour. Res. 17. 1139-1150

Brock, T. V., Lee, D. R., Janes, D. J., Winek, D. 1982. Ground water seepage as a nutrient source to a

drainage lake, Lake Medota, Wisconsin. Water research. 16: 1255-1263

Cheng, X., Anderson, M. P. 1994. Simulating the influence of lake position on groundwater fluxes. Water

Resources Research. 30(7): 2041-2049

Cherkauer, D. S. and Hensel, B. R. 1986. Groundwater flow into Lake Michigan from Wisconsin. Journal of

hydrology. 84: 261-271

Cherkauer, D. S. Zager, J. P. 1989. Groundwater interactions with a kettle-hole lake: relation of observation

to digital simulations. Journal of Hydrology. 109: 167-184.

Christensen, P. B. Sørensen, J. 1986. Temporal variation of denitrification activity in plant-covered, littoral

sediment from lake Hampen, Denmark. Applied and Environmental Microbiology. 51(6): 1174 – 1179

Downing, J. A., Peterka, J. J. 1978. Relationship of rainfall and lake groundwater seepage. Limnology and

Oceanography. 23(4): 821-825

Fellows, C. R. Brezonik, P. L. 1980. Seepage flow into Florida lakes. Water Resources Bulletin. 16(4): 635-

641


of aquatic macrophytes. Freshwater Biology. 57(5): 907-921


balance methods. Journal of hydrology. 297(1/4):187-208

Hagerthey, S. E., Kerfoot, W. C. 1998. Groundwater flow influences the biomass and nutrient ratios of

epibenthic algae in a north temperate seepage lake. Limnology and Oceanography. 43(6): 1227-1242

Hayashi, M., Rosenberry, D. O. 2002. Effects of ground water exchange on the hydrology and ecology of

surface water. Groundwater. 40(3): 309-316

Hancock, P. J., Bouton, A. J., Humphreus, W. F. 2005. Aquifers and hyporheic zones: Towards an ecological

understanding of groundwater. Hydrogeol. J. 13: 98-111

Hill, A. R., Devito, K. J., Campagnolo, S. Sanmugadas, K. 2000. Subsurface denitrification in a forest riparian

zone: interactions between hydrology and supplies of nitrate and organic carbon. Biogeochemistry. 51:

193-223

Paper 3

21


F. Ø., Pedersen, O. 2014. Role of a groundwater-Lake interface in controlling seepage of water and

nitrate. Journal of Hydrology. 517: 791-802

Kidmose J., Engesgaard P., Nilsson B., Laier T., Looms M.C. 2011. Spatial distribution of seepage to a flow

through lake: Lake Hampen, Western Denmark. Vadose Zone. Special HOBE issue. 10:110-124

Krabenhoft, D. P., Anderson, M. P., Bowser, C. J. 1990. Estimating groundwater exchange with lakes, 2.

Calibration of a three- dimensional solute transport model to a stable isotope plume. Water Resources

research. 26: 2455 – 2462

Krabenhoft, D. P., Webster, K. E. 1995. Transient hydrogeological controls on the chemistry of a seepage

lake. Water Resources Research. 31(9): 2295-2305

LaBaugh, J. W., Winter, T. C., Adomaitis, V. A., Swanson, G. A. 1987. Hydrology and chemistry of selected

prairie wetlands in the Cottonwood lake area, Stutsman County, North Dakota, 1979-1982. Us

Geological Survey. Prof.Pap., 1431, 26

Lee, D. R. 1977. A Device for measuring seepage flux in lakes and estuaris. Limnology and Oceanography.

22(1): 140-147

Lee, D.R., Cherry J.A., Pickens, J.F. 1980. Groundwater transport of a salt tracer through a sandy lakebed.

Limnol.Oceanogr. 25:45-61

Libelo, E. L. MacIntyre, W. G. 1994. Effects of surface-water movement on seepage-meter measurements of

flow through the sediment-water interface. Hydrogeology Journal. 2(4): 49-54


512

Meyboom, P. 1967. Mass transfer studies to determine the groundwater regime of permanent lakes in

hummocky moraine of western Canada. Journal of Hydrology. 5: 117-142

Ommen, D.A., Kidmose, J., Karan S., Flindt, M. R., Engesgaard, P., Nilsson, B., Andersen, F.Ø. 2012.


Denmark. Ecohydrology, 5(3): 286-296

Pfannkuch, H.O. ,Winter T.C. 1984. Effect of anisotropy and groundwater system geometry on seepage

through lakebeds. Journal of hydrology. 75: 213-237

Pfenning, K.S., McMahon, P. B. 1996. Effect of nitrate, organic carbon, and temperature on potential

denitrification rates in nitrate-rich riverbed sediments. Journal of Hydrology. 187: 283-295


Application to Lake Kinnert (Israel) and Lake Biwa (Japan). Journal of hydrology. 319(1-4): 163-175

Rittmann, B.E., Manem, J. A. 1991. Development and experimental evaluation of a steady-state,

multispecies biofilm model. Biotechnology and bioengineering. 39: 914-922

Rosenberry, D.O. (2005). Integrating seepage heterogeneity with the use of ganged seepage meters.

Limnol. Oceanogr. Methods. 3: 131-142

Rosenberry, D. O. LaBaugh, J. W. 2008. Field techniques for estimating water fluxes between surface water

and groundwater. USGS. Techniques and methods. 4-2D: 64

Paper 3

22

Rosenberry, D. O., Winter, T. C. 1997. Dynamics of water-table fluctuations in an upland between two

prairie-pothole wetlands in North Dakota. Journal of Hydrology. 191: 226-289

Sacks, L. A., Herman, J. S., Konikow, L. F., Vela, A. L. 1992. Seasonal dynamics of groundwater-lake

interactions at doñana National Park, Spain. Journal of Hydrologi. 136: 123-154

Scharling, M. 1999. Klimagrid – Danmark, Nedbør, lufttemperatur og potential fordampning 20*20 og 40*40

km. Danish Meteorological Institute. Technical report. 99-12

Sear, D. A., Armitage, P. D., Dawson, F. H. 1999. Groundwater dominated rivers. Hydrological processes.

13: 255-276

Sebestyen, S. D., Schneider, R. L. 2004. Seepage patterns pore water and aquatic macrophytes:

hydrological and biogeochemical relationship in lakes. Biogeochemistry. 68: 383-409

Shaw, R. D., Prepas, E. E. 1989. Anomalous, short-term influx of water into seepage meters. Limnology and

oceanography. 34(7):1343-1351

Shaw, R. D., Prepas, E. E. 1990. Groundwater-lake interactions II. Nearshore seepage patterns and the

contribution of ground water to lakes in central Alberta. Journal of Hydrology. 119: 121-136

Steinwand, A. L., Richardson. L. 1989. Gypsum occurrence in soils on the margin of semipermanent prairie

pothole wetlands. J. Soil. Sci. Soc. Am. 53(3): 836-842

Tóth, J. 1999. Groundwater as a geologic agent: An overview of the causes, processes, and manifestations.

Hydrogeology Jounal. 7: 1-14

Winter, T.C. 1978. Numerical simulation of steady state three-dimensional groundwater flow near lakes,

Water Resour. Res. 14: 245-254

Winter , T. C. 1983. The interaction of lakes with variability saturated porous media. Water Resour. Res. 19:

1203-1218

Winter, T. C. 1986. Effect of groundwater recharge on configuration of water table beneath sand dunes and

on seepage in lakes in the sandhills of Nebraska. Journal of Hydrology. 86: 221-237

Winter, T. C. 1999. Relation of streams, lakes, and wetlands to groundwater flow systems. Hydrogeology

Journal.7(1): 2

Paper 3

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Paper 3

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1

Paper 4

Using whole-system understanding to evaluate long

term development in alkalinity in a northern flow

through lake


2

Paper 4

1

Using whole-system understanding to evaluate long term

development in alkalinity in a northern flow through lake

Mette Frandsen*, ^, Peter Engesgaard+, Bertel Nilsson^ Ole Pedersen*

*The freshwater Biological Laboratory, Department of Biology, University of Copenhagen, Hillerød, Denmark


+Department of Geography and Geology, University of Copenhagen, Denmark

Abstract

The alkalinity in Lake Hampen, Denmark has increased from the mid-1970s until the mid-1990s. The lake

receives large amounts of nitrate from a bordering agricultural area. In this study it was examined if

denitrification of nitrate discharging to the lake from an agriculture located just north-East of the lake was of a

magnitude large enough to stoichiometrically explain the observed alkalinity development in the lake. Using

historical data on alkalinity, electrical conductivity (EC) and nitrate in the surface water and a simple mass

balance equation we estimated the historical input of nitrate to the lake. By using historical data on the EC in

the surface water we were able to use the mass balance equation to estimate the historical input of EC from

the groundwater. Nitrate is the dominant ion in the area and EC and nitrate are strongly correlated in the

ground water. Calculating the historical EC in the groundwater we could use the correlation to estimate the

historical nitrate in the groundwater and hence the historical input of nitrate via the groundwater. By

assuming the excess nitrate was denitrified, we calculated the corresponding alkalinity production.

By using this simple approach we were able to relatively accurately model the historical changes in

alkalinity. The denitrification rates we estimated are in good agreement with what other studies have found in

the lake. In conclusion, our study strongly suggests that the changes in alkalinity in Lake Hampen are mainly

caused by denitrification of nitrate discharging in large amounts to the lake from the agricultural area.

Introduction

One of the main goals in water management is to handle and find solutions in relation to obtaining and

maintaining favorable ecological status in wetlands (i.e. lakes, streams). Numerous factors threaten these

ecosystems; the more important ones on the northern hemisphere being pollution and eutrophication

(Brønmark & Hansson 2002; Brinson & Malvarez 2002). Part of the land-derived pollutants or nutrients

enters lakes and streams through groundwater seepage and understanding the interactions between

Paper 4

2

groundwater and surface water are becoming increasingly relevant in order to solve many of the challenges

posed by the Habitat and Water Framework directive (HD and WFD).

At catchment scale groundwater-surface water systems are complex united water bodies of

interrelated hydrological, biological, and chemical processes. In order to understand the impact on surface

water from the surrounding catchment, data on different scales and related to different disciplines are

required (Zalewski, 2002). These interrelationships are complex and the lack of sufficient spatial and

temporal data to estimate the total in- and outputs are often not available and, if available, they often

represent point scale measurements (Moss, 1999; Rodriguez-Iturbe, 2000). Up-scaling point observations to

whole system analysis are often succumbed to high uncertainty, and tend to be unique to the given system

they describe. Hence, one of the major challenges in water management is one of scale in space and time

and data availability.

One way to overcome these challenges is to use historical data and try to explore temporal trends

and couple these trends to the overall in- and outputs of the whole system. In freshwater lakes, the changes

in water quality can be seen as an integrated response to all the catchment specific in- and outputs as well

as the in-lake processes (Krabbenhoft & Webster, 1995; Gurrieri & Furniss, 2004; Rimmer et al., 2006).

Hereby, the lakes acts as a sentinel, tracking changes in the catchment (Williamson et al. 2009) and they are

able to retain a solute influx memory storing information about the lake-watershed-climate relationship in the

past (Rimmer et al. 2006). This memory is available in the form of chemical surface water data containing

more information about the overall in- and outputs of a system, than point scale measurements. In the

surface water, seasonal mixing ensures an overall representative measure for the net effect of all in- and

outputs. In groundwater, little mixing occurs and extrapolation of data collected from point measurements

can be highly uncertain. If the hydrology of the system is understood with the in- and outputs of water known

for a given system, then time series of solute concentrations in the surface water can be used as proxy for

estimating the net loading of solutes to the surface water (Rimmer et al. 2006).

Both an increase in nutrient addition to a lake and the following increase in primary production will

affect the alkalinity of the system (Davison 1986). Eutrophication and internal alkalinity production is

interlinked (Sigg et al., 1991; Davidson, 1986; Schindler et al. 1986) as the alkalinity production is driven by

reduction of the major anions such as sulfate and nitrate (Schindler 1986; Rudd et al. 1986). In soft water

lakes these processes can be significant factors in regulating the acid-base system (Davidson, 1986). As a

consequence, the alkalinity in many soft water lakes has increased as an effect of increased nutrient loading

during the past 50 years (Sutcliffe et al. 1982).

When investigations of alkalinity changes of freshwater lakes started in the early 1930s, it was

predominantly assumed that alkalinity production primarily derived from ion exchange and weathering of

terrestrial soils and rocks (Schindler, 1986). From the late 1970s, in-lake processes such as sulfate reduction

(Hongve, 1878; Schindler et al. 1980; Cook et al. 1986), biological productivity (Kelly et al. 1987), and

denitrification (Davidson 1986, Schindler, 1986) were included. In 1986, Schindler et al. found that in-lake

alkalinity was higher than in the inflowing streams, pointing towards the importance of internal sources of

alkalinity. They showed that more than half of the in-lake alkalinity production was biological rather than

Paper 4

3

geochemical. The major alkalinity generating processes were the reduction of SO42-

, exchange of H+ for Ca

2+

in sediments and biological reduction of NO3- (denitrification).

More recently it has been acknowledged that when investigating alkalinity changes over long time

scales, historical factors may be important. With the introduction of the combustion engine and the electrical

power generation in the 1900th century, acid rain from oxidation of fossil fuels, increased in most parts of

Europe (Schindler 1988). In Denmark, sulfate concentration in precipitation increased by 50% from the late

1950s to the early 1970s (Rohde & Rood, 1984) having a significant effect on the alkalinity of some inland

lakes (i.e. sulfate reduction).

Lakes connected to cultivated land through groundwater flow can be subjected to large amounts of

nitrate input. The anaerobic denitrifiers need nitrogen in the form of NO3-(nitrate) and this is often supplied

from the aerobic nitrifying processes where NH4+ (ammonium) is oxidized to NO2

-/NO3

- (Appelo & Postma,

2005). However, if the nitrogen is readily supplied in the form of nitrate, the denitrifiers only need a sufficient

electron donor (i.e. carbon, pyrite or Fe2+

) to drive denitrification. Hence, in lakes with cultivated catchments,

the alkalinity generation through denitrification can play a significant role in the alkalinity generation in a lake.

Alkalinity generation coupled to denitrification primarily takes place at the sediment-water-interface

(SWI) under anoxic conditions (Appelo & Postma 2005).

6NO3- + 5 CH3OH 3N2 + 5CO2 + 7H2O + 6OH

-

It follows that one mole of alkalinity is produced, when one mole of nitrate is removed. As long as there is a

sufficient pool of inorganic carbon the denitrification rates are coupled to the accessibility to nitrate. This

process can be very distinct in space. Hill et al. (2000) show that denitrification hotspots occur in terrestrial

areas where nitrogen in the form of nitrate is available through groundwater. The same can be the case near

lakes, where nitrate from surrounding cultivated land is transported to the SWI. Hence, in lakes that receive

large amounts of nitrate-rich water, the alkalinity of the system can be greatly affected.

In this study we examined if historical nitrate input to Lake Hampen, Denmark, can explain the

observed increase in alkalinity from the early 1970s until the mid-1990s by using a simple mass balance

approach and data on end members of water entering the lake. Previous studies have found that the

agricultural input of nitrate accounts for up to 67% of the total nitrogen input to the lake, even though it only

account for 23% of the groundwater discharge to the system (Kidmose et al. 2014).

As a result it is shown that changes in nitrate input can lead to changes in the denitrification

processes at the SWI, and that the magnitude of these changes can account for the alkalinity increase

observed in the system.

Paper 4

4

Method

Study site

Lake Hampen is located in the central part of Denmark just west of the Jutland ridge and just east of the last

main glacial advance. The lake is situated in an area consisting mainly of coarse melt water sands and

gravel reaching ~25m down below the surface. The catchment (993 ha) is primarily covered with forest

(62%) and agricultural land (30%) (Figure 1). The agricultural land is located on the north-eastern shoreline

where most of the groundwater discharges to the lake. Lake Hampen is an oligo- to mesotrophic soft water

lake with a surface area of 76 ha, and a maximum and mean depth of 13 and 4m respectively.

Hydraulically, the lake is characterized as a seepage lake receiving 2/3 of its water from

groundwater and the rest from precipitation (Table 1). The directional flow of groundwater is from the north-

east, east and south-east where the groundwater discharges from the aquifer to the lake and to the West,

where it recharges the aquifer from the lake (Figure 1). Since 1977, the lake has been a habitat reference

area, but the lake has undergone continuous eutrophication dating back several decades. In the early 1980s

an illegal spillage of manure was discovered and stopped. On top of that, large amounts of nitrate discharges

to the lake from the bordering agricultural site, where the cropped field almost reach the water line of the lake

(Kidmose et al., 2011; Ommen et al., 2012; Karan et al., 2014;).

Hydrological data and groundwater-lake exchanges

Precipitation (as annual averages) originates from the Danish Metrological Institute Danish Meteorological

Institute, (Frich et al. 1997) (Table 1, Figure 2). The water chemistry for electrical conductivity (EC) in

precipitation is taken from Uglebjerg (2013, unpublished data.) and nitrate from Kidmose et al. (2011).

The water budget for 2008 was estimated by Ommen et al. (2012) on the basis of a lake segment

approach using well information and Darcy. Groundwater seepage fractions was estimated by Knudsen 2013

Figure 1). Lake Hampen showing locations of wells and simulated potentiometric lines (meters above sea level) modified from Kidmose et al. (2011). The catchment is primarily covered with conifer forest, but on the North-Eastern shoreline cropped fields borders the lake. The location of wells W1-W4 and DX6 are shown.

Paper 4

5

(unpublished data ) using transient data collected by Kidmose et al. (2011) and data collected during this

study (2010-2012). The transient groundwater-lake system was simulated by modifying the model of

Kidmose et al. (2011) using the module LAK3 in Modflow (Merritt & Konikow, 2000). The results show that

the system generally can be regarded as at quasi steady-state.

Lake and groundwater chemistry

To monitor the system response to nitrate input from agriculture, historical data from 1971 - today were used.

Data from 1971 to 1998 was obtained from a national monitoring program (Moeslund, 2000). Data are used

here as average values from March to September. No winter data are available. Data from 1998 to 2010

were collected from various studies conducted on the lake in the past 12 years (not published).

Data on groundwater chemistry was collected from five deep wells; W1 (screened to 16 m below

land surface), W2 (screened to 14 m below lake bed), W3 (screened to 6.4 m below lake bed), W4 (screened

to 14 m below land surface) and DX6 (screened to 17 meter below land surface). The wells represent

groundwater discharge from the agricultural site (W1, W2, W3 and DX6) and groundwater from areas not

affected by agriculture, but from mixed forest (W4). Wells W2 and W3 were installed 7 and 22 m off-shore,

respectively. The wells were constructed of galvanized steel pipes with an outer diameter of 0.025 m and a

screen length of 9 cm. Water samples were collected after clean pumping three times for every meter and

EC was measured. Water samples for chemical analysis were stored at 5 oC until analysis. Dissolved nitrate

was analysed spectrophotometrically on an automated ion flow injection analyser (QuickChem methods 10-

107-04-1-C) (McKnight, 1986). In W4 only EC was measured.

Mass balance approach

A four-step mass balance approach was used to calculate the historical changes in lake alkalinity. The lake

is assumed to be at steady state with respect to flow (Knudsen, 2013). A mass balance equation adopted

from Krabbenhoft & Webster (1995) is used in two of the steps;

(1)0VkCSCGCSCGPCdt

dM0LoLoiiGiP

L

where dML/dt is the change in mass in the lake over time, the first three terms on the right hand side are the

inputs and the last two terms are the outputs. P is the total precipitation onto the lake surface (m3/year), Gi

and Go the total groundwater discharge to and recharge from the lake (m3/year), and Si and So surface water

inflows and outflows (m3/year). Note that evaporation is not included. The lake is assumed completely mixed

with concentration CL. Concentrations are specified for all fluxes, i.e., CP, CG, and Ci for precipitation,

groundwater, and inlet surface water. Note that all outflows are given the lake concentration. A zero-order

mass removal term is added, k0, to account for any in-lake processes in the total volume of the lake, V. The

calculations were performed with yearly time steps and since the residence time of water in the lake is

approximately 1.5 years (Ommen et al., 2012) a steady-state assumption is adopted, i.e., dML/dt=0.

Paper 4

6

The Gi term was split into two components since the groundwater input originates from both

agricultural and forest areas, i.e. Gia and Gif, respectively with concentrations CGa and CGf, respectively. The

surface water inlet (Si) is negligible (Kidmose et al., 2011) and the surface water outlet is small and the two

outflows are therefore combined into one Got (=Go+So). With these simplifications the mass balance reads;

(2)0VkCGCGCGPC 0LotGfifGaiaP

Because it is assumed that all flow inputs and outputs balance one has Got=(P+Gia+Gif)=Git, where Git is the

total input of water. By dividing (2) with either Got or Git (2) reads;

(3)0τkCFCFCFCF 0LOGfFGaAPP

where Fi are the fractions of inflows or outflow from precipitation or groundwater. is the residence time

(V/Got). The fractions Fi are known from the 3D groundwater-lake simulations carried out by Kidmose et al.

(2011), see above and Table 1, and are all nearly constant in time (Knudsen, 2013, not published). The

concentrations of groundwater originating from forest areas and in precipitation were assumed known and

constant in time based on data supplied by our investigations and the works of Ommen et al. (2012) and

Uglebjerg (2013, unpublished data), as discussed above. This leaves CGa(t), CL(t), and k0 as unknowns (here

it is assumed that =1.5 years). Note that the concentrations in groundwater and the lake vary in time.

Step 1: Historical EC in groundwater discharging from agricultural area

Historical concentrations of surface water EC was used to predict the concentrations of EC in groundwater

discharging from groundwater using (3);

(4)F

τkCFCF -(t)CF)t(C

A

0LPPGfFLOGa

where all units are in S cm-1

. The term k0 was fitted to the early curve of historical observations to account

for any in-lake processes changing the water chemistry and thus either removing or producing EC. The

background EC in groundwater from forested areas was evaluated from well W4.

Step 2: Historical nitrate concentrations in groundwater discharging from agricultural area

Data from wells W1 to W3 + DX6 were used to correlate measured EC and nitrate in groundwater

discharging from the agricultural area. Hereby it is possible to calculate the historical nitrate concentrations

using the correlation structure and the predicted historical concentrations in EC as a result of step

Paper 4

7

Step 3: Historical nitrate concentrations in surface water

Equation (3) was used to predict the historical concentrations of nitrate in the lake, i.e.;

(4)F

τkCFCF (t)CF)t(C

O

0LPPGfFGaAL

where all concentrations are given in mmol L-1

. The term k0 is set equal to zero and the difference between

observed and calculated nitrate concentrations was evaluated, i.e., NO3. This residual was regarded as the

time-varying nitrate removal.

Step 4: Historical alkalinity in surface water

Any difference between calculated and observed nitrate concentrations in the surface water (NO3) was

assumed to be caused by denitrification at the SWI, i.e., alkalinity is produced at the lake bed and is a source

of alkalinity to the lake. As one mole of alkalinity is produced per mole of nitrate reduced the alkalinity can be

estimated as;

(5)NOALK 3

In a final step 5 it was attempted to quantify denitrification rates at the SWI.

Step 5 Estimating denitrification at the SWI

The main source of nitrate to the lake is from the agricultural area (Ommen et al., 2012; Kidmose et al.,

2014). If assuming that nitrate is completely removed through denitrification (producing alkalinity), the rate of

removal of nitrate per m2 lake bed is calculated as;

(6)(t)CQ

)(r GaGad

At

where rd(t) is the annual denitrification rate (mole/(m2*hour)) over the lake bed area (A). The area A was

provided by the groundwater-lake models of Kidmose et al. (2011) and Knudsen (2013, not published). The

denitrification rate was compared to previous in-situ measurements and literature values.

Paper 4

8

Results

Water budget

Lake Hampen receives ~2/3 of its water from

the groundwater (Kidmose et al. 2011). Knudsen

2013 (unpublished data) estimates that 14% of

the groundwater input comes from the

agricultural catchment, 52% comes from the

forest catchment, and the remaining enters the

lake through precipitation. Of the total volume of

incoming water 75% leaves the lake through

groundwater recharge and the rest through

evaporation (Table 1).

Historical development in alkalinity, EC and

nitrate

The long term alkalinity development in Lake

Hampen shows a clear increase from 1980 until

the mid-1990s (Figure 2, c). In mid 1970s the

alkalinity was relatively stable around 0.1-0.2

mmol L-1

. From around 1980 the alkalinity

started to increase reaching ~ 0.5-0.6 mmol L-1

in the 1990s. In 1995 a peak in alkalinity was

observed reaching the highest concentration

ever recorded in the lake with an alkalinity of

0.59 mmol L-1

. After the peak in 1995, the

concentration stabilized on the levels observed

in the 1980s. In 2008-2010, the alkalinity was

around 0.3 mmol L-1 (Figure 2, c).

The EC followed an almost congruent

pattern (Figure 2b). The EC was generally high

from the mid-1980s to the mid-1990s. In 1995, there was a peak where EC reached the highest recorded EC

in the lake of 249 S cm-1

. This is not directly apparent from the figure as this only shows the average

concentration from May-September for each year.

Figure 2) Figure shows a) precipitation from 1970 to 2013; b) historical development in surface water EC, c) alkalinity and d) nitrate. Data presented for EC, alkalinity and nitrate are average concentrations from May to September for each year when available (see Moeslund, 2000).

Paper 4

9

After the peak in 1995, EC stabilized again on a slightly higher level than prior to the 1995 peak.

The nitrate concentration followed a less distinct pattern (Figure 2, d). During the 1970s the

concentration fluctuated between 0.003 and 0.01 mmol L-1

, averaging 0.006 mmol L-1

. During the 1980s

therewas little data, but in 1995 a peak in nitrate concentration was observed (as seen in alkalinity and EC)

with concentrations reaching 0.03 mmol L-1

. In the years after the 1995 event, there was again little data, but

the concentration leveled out on a lower level than prior to the event.

Precipitation

The average precipitation over the whole period was 856mm/year with a minimum of 521 mm in 1996 and a

maximum of 1129 mm in 1981 (Figure 2, a). The EC in the precipitation was estimated using data from 2011

(Uglebjerg, 2013, unpublished data). The average EC was 54 µS cm-1

and is assumed representative for the

whole period. The nitrate concentration in precipitation, 0.01 mmol L-1

, was estimated using data from

Kidmose et al. 2014

Groundwater chemistry

EC and nitrate in groundwater in wells W1-W3 and DX6 are strongly correlated (Figure 3). In W1 there is a

strong positive correlation between nitrate and EC (R2=0.55, P<0.005, Pearson) (Figure 3). The nitrate

concentration was highest 6 m below the surface reaching 80 mg L-1

. Below this the concentration declines

to ~10 mg L-1

and was relatively stable from 10-14 m. The EC was highest at 3-5 m depth reaching almost

400 µS cm-1

. Below this depth, EC decreased slowly from ~400 to ~ 200 µS cm-1

at 14 m depth. In W2 (a well

approximately 10 m off-shore) EC and nitrate were measured for every 0.50 m, and the pattern was the

same as in W1. There is a strong positive correlation between nitrate and EC (R2=0.73, P<0.0001, Pearson)

(Figure 3). The nitrate concentrations in the top 4 m of the sediment below the lake bed varied around 75 mg

Figure 3. Development in EC (µS cm-1

) and nitrate (mmol L-1

) with depth in wells W1, W2, W3 and DX6.

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10

L-1

but with a drop to 50 mg L-1

at 2.5 m depth. From 4 – 7 m depth

the nitrate concentrations were relatively stable around 80 mg L-1

and below the concentration declined steadily to around 10 mg L-1

until 10 m depth. In 10-14 m, the concentration was relatively stable

around 10 mg L-1

. The EC followed the same pattern. However, at

5 m depth, where there was a sudden decrease in nitrate, the EC

exhibited a sudden increase from 277 µS cm-1

at 1m depth to 436

µS cm-1

at 2.5 m depth. From 4-7 m depth, EC decreased and

increased again from ~340 µS cm-1

with a minimum at 5 m depth

with an EC of 274 µS cm-1

. Hereafter EC decreased to 274 µS cm-1

at 5 m depth to 220 µS cm-1

in 14 m depth.

In W3 (a well 22 m off-shore), there was a strong positive

correlation between nitrate and EC (R=0.68, P>0.0005, Spearman)

(Figure 3). The nitrate concentration was very low in the top 2 m

with concentrations of only 0.35 mgL-1

at 0.40 m depth. From 0.40

to 4.40 m depths the concentration increased to 90 mg L-1

with a

drop at 3.40 m depth to ~45 mg L-1

. From 4.40 m and down to 6.40

m depth the concentration declined slowly from 90 mg L-1

to ~80

mg L-1

. EC increased from 100 to ~500 µS cm-1

in the first 4 m.

Below, the EC declined reaching ~300 µS cm-1

at 6.40 m depth.

In DX6 there was a strong positive correlation between

nitrate and EC (R2=0.73, P<0.0005, Pearson). The nitrate concentrations decreased slowly from 100 mg L

-1

at 2m depth to 90 mg L-1

at 8 m depth followed by a steeper decrease from 8-13 m depth from ~90 to ~9 mg

L-1

. The EC was high in the top 3 m reaching 550 µS cm-1

declining to ~280 µS cm-1

at 5 m depth. From 5 – 8

m depth EC increased again to 420 µS cm-1

and after that decreased slowly to ~220 µS cm-1

in 13m.

Figure 4. EC (µ cm-1) in well W4

sampling groundwater from

forested area. EC in this area is

not affected by nitrate and thus is

used as a background EC for the

area.

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11

In W4 only the EC was measured. The EC was stable around 200 µS cm-1

at all depths (figure 4).

The EC measured in this well was unaffected by nitrate and this value was used as a background EC for the

groundwater discharging from forested areas.

Mass balance model

Step 1: Historical groundwater EC and Nitrate concentrations on the agricultural site

Solving (1) for CGA we get a historical estimate of the EC at the agricultural site (Figure 5). A k0 of 50 µS cm-1

year-1

was fitted to match the early part of the curve, where the influence of high nitrate concentrations were

small as these generally are delayed by 10 years (see also below).

The estimated EC in groundwater at the agricultural discharge area was approximately 2-4 times

higher than the EC observed in the lake. In comparison, the other end-members were 200 and 54 µS cm-1

for

groundwater discharging from forested areas and in precipitation, respectively. According to this the majority

of the ions enter the lake through groundwater discharge from the agricultural area, which only accounts for

14% of the total input of water.

Step 2: Historical nitrate concentrations in groundwater discharging from agricultural area

The estimated EC in groundwater was used to estimate the historical concentrations of nitrate in

groundwater at the agricultural transect. A correlation analysis between EC and nitrate was conducted on all

groundwater data from the agricultural site (W1-W3 and DX6). A background value of 200 µS cm-1

was first

subtracted from the EC data. The 200 µS cm-1

is the baseline EC in the area where no nitrate is present

(data from well W4 in forest, Figure 4).

There is a good correlation between EC and nitrate (p<0.001, R2=0.48, Pearson) (Figure 6). Using

this correlation, we estimated nitrate in groundwater at the agricultural transect based on the estimated EC

concentrations.

Figure 5) Observed EC (µ cm

-1) in the surface water

(solid line) and estimated EC in the seepage discharge at the agricultural transect (stippled line).

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12

The concentration of nitrate at the

agricultural site was very high and greatly

exceeded the surface water concentration

(Figure 7). Naturally, the shape of the

curve followed that of EC in groundwater.

The concentration increased steadily from

the start to the mid-1990s. In the mid-

1990s there was a large peak, both in the

observed concentration in the surface

water and in the calculated groundwater

concentration. The groundwater

concentration reached 3.2 mmol L-1

which

is more than 160 times higher compared to

the observed peak in the surface water. After the mid-1990s the concentration decreased again to levels

slightly lower than before the peak.

Step 3: Prediction of historical nitrate concentrations in the lake

Finally, the mass balance equation was used to solve for Clake (nitrate) to obtain a historical prediction of

surface water concentration of nitrate without any removal processes and with the estimated input from the

groundwater from step 2.

The modeled nitrate concentration in the lake was from 10-200 times higher than the concentration observed

in the lake (Figure 8). The average, minimum, and maximum concentration of nitrate in the lake during the 40

year period were 0.21, 0.04 and 0.6 mmol L-1

, respectively compared to the observed concentrations of

0.007, 0.0007 and 0.03 mmol L-1

, respectively. The discrepancy between modeled and observed

Figure 6) Correlation (Pearson correlation) of NO3 and EC on data from W1, W2, W3, and DX6.The solid line shows the linear regression for NO3 and EC from all wells. The grey dotted line shows the regression line for all pairs except where EC is below 200 µS cm.

Figure 7) Calculated nitrate in the groundwater at the agricultural transect (NO3

-a) (left axis) and observed

nitrate in the surface water (NO3-L) (right axis).

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13

concentration was greatest around the 1995 peak

Step 4: Estimated alkalinity production through denitrification

Denitrification of nitrate was stoichiometrically balanced with the amounts of alkalinity produced in a 1:1

relationship (see Materials and Methods). If we assume all the excess nitrate (difference between the two

curves in figure 7), is denitrified, the alkalinity of the lake would be as shown in Figure 9. This is surprisingly

close to the observed historical changes in alkalinity in the lake.

Step 5 Estimating denitrification at the SWI

Nitrate primarily enters the lake through groundwater seepage at the agricultural catchment. The discharge

area in the lake bordering this transects is estimated to be ~14.000 m2 (Knudsen, 2013, unpublished). To

further test our assumption of a denitrification induced alkalinity increase in the lake, we estimated the

denitrification rates required to denitrify all the nitrate coming from the agricultural transect using eq. (6). The

calculated denitrification rates vary between 10-130 µmol N h-1

m-2

(Figure 10).

Discussion and conclusion

Using long time series of data on surface water chemistry dating back to the 1970s and a simple mass

balance approach we show that denitrification of nitrate-polluted groundwater coming from the agricultural

part of the catchment can be an explanation for the increase in alkalinity observed in the lake during the 80s

and 90s.

Figure 9) Modeled and observed

alkalinity. The relationship between

observed and modeled alkalinity was

significant (Pearson correlation, P<0.005)

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14

The historical development in

alkalinity from 1971 till today shows a

marked increase starting in the mid-1980s

peaking in 1995 hereafter decreasing

slowly. During this period, the precipitation

has been fairly constant and we found no

correlation between precipitation and

alkalinity when looking solely on the

volume of rain. However, in 1995 the

precipitation was low, which could to some

extent explain the large peak observed for

both alkalinity and nitrate.

Using the mass balance equations and the correlation between nitrate and EC in groundwater, we

estimated the historical nitrate concentrations in groundwater discharging from agriculture. The estimated

nitrate concentrations are in agreement with recent measurements from the area. We estimate

concentrations reaching up to 1.5 -1.7 mmol L-1

with a peak in 1995 of 3.2 mmol L-1

. In comparison, Kidmose

et al. (2014) found nitrate concentration reaching 2.8 mmol L-1

in the same area.

Our results are in agreement with a national trend linking an increase and subsequent decrease in

groundwater nitrate concentrations to regulation of agricultural practices. During the past 100 years, the

agricultural production in Denmark has increased significantly. This was reflected in the nitrate

concentrations in the Danish groundwater and a clear increasing trend was observed following the increased

use of fertilizers especially since the 1950s (Hansen et al. 2012). In the mid-1980s the government imposed

regulations, demanding that N leaching from agricultural practices should be reduced (the NPO regulation,

Danish ministry of the environment, 1984).

Hansen et al. (2012) have conducted

a large survey analyzing the national trends in

nitrate in the groundwater before and after the

NPO regulation and they clearly show these

trends. From the 1970s until the NPO

regulation, the nitrate concentrations in shallow

oxic groundwater increased, after which a

slower decline since 1985 have been observed.

This is the same pattern we found in

this local investigation. We estimate an

increasing nitrate load to Lake Hampen from

around 1980 until the 1995 peak followed by a

decrease (Figure 8). In our study, we have a

ten-year delay compared to the national trend.

Figure 8) calculated nitrate in the lake (NO3- L

mod) and observed nitrate in the lake (NO3- L).

1970

1975

1980

1985

1990

1995

2000

2005

2010

0

50

100

150

200

250

Year

Denitri

fication (m

ol N

m-2

h-1

)

Figure 10) Estimated denitrification rates if all excess nitrate is denitrified in the area that discharges nitrate rich water from the agricultural catchment.

Paper 4

15

Kidmose et al. (2011), however, show that groundwater discharging from the agricultural fields at Lake

Hampen on the average is 10 years old, so a delay of 10 years can be expected.

Using our model we estimated the nitrate concentration in the discharging groundwater to be

around 1.08 mmol L-1

(average for the whole period). Only a small fraction of this can be recovered in the

lake. This means that nitrate in the discharging groundwater from the agricultural transect should be either

denitrified or lost via other processes before entering the lake.

To denitrify all, nitrate reaching the lake via the groundwater in our model, denitrification rates

reaching 10-130 µmol N h-1

m-2

(with a peak of 214 µmol N h-1

m-2

in 1995) would be needed (Figure 10).

Typically the denitrification rates in freshwater lake sediment are in the range of 2-171 µmol N h-1

m-2

normally not surpassing ~60 µmol N h-1

m-2

(Seitzinger, 1988). However the variability between systems is

quite large. Christensen & Sørensen (1986) measured denitrification rates on sediment cores form Lake

Hampen. They found denitrification rates of approximately 50 µmol h-1

m-2

(summer values). However, when

supplying nitrate, they saw a 7-fold increase in denitrification activity reaching rates of 225 – 350 µmol N h-

1m

-2. Therefore, it seems that the sediments can support higher denitrification rates and are limited by nitrate.

In the specific area of Lake Hampen we examined, there is a continuous supply of nitrate in high

concentrations, and our estimated rates seem realistic.

The nitrate-rich discharge supplies a ready-to-use electron acceptor. Normally nitrate is fed to the

denitrification from nitrification of ammonia (Lohse et al. 1993) making the process dependent on both the

nitrification processes and of varying periods of oxic (nitrification) and anoxia (denitrification) conditions. This

could significantly lower denitrification rates. Besides nitrate the denitrifying bacteria need a carbon source.

In lake Hampen, we have measured high concentrations of organic matter in the sediment on the agricultural

transect reaching up to 2.4 % (not published, own data). Furthermore, the often significant release of organic

substances from the plant roots, as shown by Søndergaard (1983), also feed the denitrification processes.

Based on our results as well as the results from these previous studies we find that our suggested

denitrification rates are well within the boundaries of what would be possible in a system like this.

When denitrification is uncoupled from nitrification, it can have large impact on the alkalinity.

Denitrification is an alkalinity producing process whereas nitrification is an alkalinity consuming process.

These two processes tend to equal out each other (Davison 1986; Risgaard-Petersen & Jensen 1997), but in

systems where nitrate is readably supplied denitrification possibly outruns the nitrification and a surplus of

alkalinity is produced.

Numerous studies suggest that denitrification could be a major factor in increased alkalinity. Baker

et al. (1988) showed that in-lake alkalinity production was important in soft water lakes. Carignan (1985) that

alkalinity production in the sediment of lakes could neutralize almost all acid inputs to the system. Rudd et al.

(1988) showed that denitrification is an important source of alkalinity in soft water lakes. Abril & Frankignoulle

(2000) showed that bacterial processing of nitrogen both rapidly and strongly affects the alkalinity of the

system.

The alkalinity production leads to an alkalinity increase in the water of Lake Hampen fitting the

observed alkalinity well. The modeled data is slightly higher compared to the observed data, and the peak in

1995 is overestimated. This could be caused by the somewhat simplified approach we use. We possibly

Paper 4

16

overestimate the EC in the groundwater from the agriculture and this leads to an overestimation of nitrate

from the agriculture followed by and overestimation of alkalinity production. The EC in the groundwater is

posibly overestimated in the mid-1990s because we use the lake concentration of EC to calculate the input

from the different sources. In the mid-1990s we see the lowest precipitation in the whole period. This would

lead to higher surface water EC as the lake stage, the seepage rates and the water renewal times would all

be lower (Schindler, 1986). Hence, in our model this high EC leads to an overestimation of EC in the

groundwater even though this is not the case in the real situation.

We also assumed that the EC in the rain and from the groundwater discharging from the

forest areas have been constant over the considered period. It could be speculated that the lake have been

subject to acidification from precipitation and that the alkalinity increase is a recovery from this atmospheric

acidification. We have no way of testing this, but, on the other hand, the alkalinity increase follows nicely the

national development in nitrate concentrations in groundwater. Furthermore, the 1995 event cannot be

explained with a slow recovery from an atmospheric input but points more towards our theory of a

denitrification-induced alkalinity production.

Alkalinity increase can have substantial impact on the ecosystem and the distribution and

community composition of the submerged vegetation (Marberly & Spence 1983). Vestergaard & Sand-

Jensen (2000) showed that alkalinity is a main factor responsible for species distribution of submerged plants

in Danish soft water lakes and represents a threat to unique isoetid communities. In low alkaline waters,

availability of inorganic carbon can be very limiting for growth. Here slow growing isoetids species dominate

as they are morphologically adapted to access the pool of inorganic carbon in the sediment (Sand-Jensen &

Prahl, 1982; Madsen et al. 2002). The isoetids are poor competitors to the fast growing elodeid species, but

their adaption to sediments scavenging for both nutrients and inorganic carbon enable them to thrive in water

not suitable for the elodeids (Maberly & Madsen, 2002). With increased alkalinity the availability of inorganic

carbon, especially in the form of HCO3- increases. This favors the elodeids, and increases the competition

with the isoetids.

With this study we show that leaching of nitrate from agriculture can have a substantial impact

on the lake chemistry. Not only does the nitrate affect the lake directly by adding nutrients, but also indirectly

by supplying nitrate for the denitrification processes that ultimately leads to increased alkalinity in the lake. In

a historical time frame, we were able to stoichiometrically link the denitrification of nitrate with the overall

alkalinity development in the lake hereby stressing the importance of both in-lake processes as well as the

groundwater dependent input to the lake. By using a simple mass balance equation we were able to fit the

modeled alkalinity production with the observed alkalinity development in the lake over a period exceeding

30 years. Long term measurements of surface water quality are therefore important to track changes in land

use and hydrology in the catchment.

Acknowledgements

This project was founded by The Danish Council for Independent Research – Nature and Universe.

Paper 4

17

References

Abril, G. and Frankignoulle, M. 2000. Nitrogen-alkalinity interactions in the highly polluted Scheldt basin

(Belgium). Water resources. 35(3): 844-850

Appelo, C.A. J. & Postma, D. Geochemistry, Groundwater and pollution. 2005. A.A. Balkema publishers.

458-462.

Baker, L. A and Brezonik, P. L. 1988. Dynamic model of in-lake alkalinity generation. Water Resources

research. 24(1): 65-74

Brinson, M. M., Malvarez, A. I. 2002. Temperate freshwater wetlands: types, status, and threats.

Environmental Conservation. 29(2): 115-133.

Brønmark, C., Hansson, L. 2002. Environmental issues in lakes and ponds. current state and perspectives.

Environmental Conservation. 29(3): 290-306

Carignan, R. 1985. Quantitative importance of alkalinity flux from the sediment of acid lakes. Nature. 31: 158

– 160

Christensen, P. B., Sørensen, J. 1986. Temporal variation of denitrification activity in plant-covered, littoral

sediment from Lake Hampen, Denmark. Applied and environmental microbiology. 51(6): 1174 - 1179

Cook, C. A., Schindler, D. W. and Turner, M. A. 1986. Mechanisms of hydrogen ion neutralization in an

experimentally acidified lake. Limnology and Oceanography. 31(1): 134-148.

Davison, W. 1986 Internal element cycles affecting the long-term alkalinity status of lakes: implications for

lake restoration. Scheiz. Z. Hydrol. 49(2): 186-201

Frich, P., Rosenørn, S., Madsen, H. and Jensen, J.J. 1997. Observed precipitation in Denmark, 1961-90.

Technical Report 97-8. Danish Meteorological Institute.


balance methods. Journal of Hydrology. 297: 187-208.

Hansen, B., Dalgaard, T., Thorling L., Sørensen, B., Erlandersen M. 2012. Regional analysis of groundwater

nitrate concentrations and trends in Denmark in regard to agricultural influence. Biogeosciences, 9,

3277-3286

Hill, A. R., Devito, K, J., Campagnolo, S. & Sanmugadas, K. 2000. Subsurface denitrification in a forest

riparian zone: Interactions between hydrology and supplies of nitrate and organic carbon.

Biogeochemistry. 51: 193-22

Hongve, D. 1978. Buffering of acid lakes by sediments. Verh. Int. Ver. Limnol. 20: 743-748.


F. Ø., Pedersen, O. 2014. Role of a groundwater-Lake interface in Controlling seepage of water and

Nitrate. Journal of Hydrology. 517: 791-802

Kelly, C. A. Hesslein, R. H., Shindler, D. W., Dillon, P.J., Driscoll, C.T., Gherini, S.A. and Hecky, R.E. 1987.

Prediction of biological acid neutralization in acid-sensitive lakes. Biogeochemistry. 3: 129-140

Kidmose, J., Engesgaard, P., Nilsson, B., Laier, T. and Looms, M. C. 2011. Spatial distribution of seepage at

a flow-through lake: Lake Hampen, eastern Denmark. Vadose Zone Journal. 10(1): 110-124

Paper 4

18

Kidmose, J., Engesgaard, P., Ommen, D. A., Nilsson, B., Flindt, M. R., Andersen, F. Ø. 2014. The role of

groundwater for lake-water quality and quantification of N seepage. Groundwater.

DOI10.1111/gwat.12281

Knudsen, K. F. 2013. Effect of climate on groundwater exchange to and from Lake Hampen – a 3D

groundwater lake simulation study. Master thesis. Department of Geosciences and Natural Resource

Management. University of Copenhagen.

Krabbenhoft, D. P., Webster, K. E. 1995. Transient hydrological controls on the chemistry of a seepage lake.

Water Resources research. 31(9): 2295-2305

Lohse, L., Malschaert, J. F.P., Slomp, C. P., Helder, W. and Raahorst, W. V. 1993. Nitrogen cycling in North

sea sediments: interaction of denitrification and nitrification in offshore and coastal areas. Marine

ecology progress series. 101: 283-296

Marberly, S.C. and Madsen, T.V. 2002. Freshwater angiosperm carbon concentration mechanisms:

processes and patterns. Functional Plant Biology. 29: 393-405

Marberly, S.C. and Spence, D.H.N. 1983. Photosynthetic inorganic carbon use by freshwater plants. Jounal

of Ecology. 71: 705-724

McKnight, R. 1986. Quichem method No. 10-107.04-1-c. Lachat Chemicals inc. Mequon.

Madsen, T.V., Olesen B. and Bagger, J. 2002. Carbon acquisition and carbon dynamics by aquatic isoetids.

Aquatic Botany. 73: 351-371

Merritt, M. L. & Konikov, L. F. 2000. Documentation of a computer program to simulate lake-aquifer

interaction using the MODFLOW ground-water-flow model and the MOC3D solute-transport model,

USGS. Water-resources investigation report. Series number 2000-4167

Moeslund, B. 2000. Lake Hampen: Environmental status 1971-1999. Vejle County (in Danish)

Moss, B. 1999. Ecological challenges for lake management. Hydrobiologia. 395(396): 3-11

Niswonger, R.G. & Prudic, D. E. 2005. Documentation of the streamflow-routing (SFR2) package to include

unsaturated flow beneath streams – a modification to SFR1. USGS techniques and methods. 6/A13: 50

Ommen. D. A. O., Kidmose, J., Karan, S., Flint, M. R., Engesgaard, p., Nilsson, B. and Andersen, F. Ø.

2012. Importance og groundwater and macrophytes for the nutrient balance at oligotrophic Lake

Hampen, Danmark. Ecohydrology. 5(3): 286-296


Application to Lake Kinneret (Israel) and lake Biwa (Japan). Journal of Hydrology. 319: 163-175

Risgaard-Petersen, N. and Jensen, K. 1997. Nitrification and denitrification in the rhizosphere of the aquatic

macrophyte Lobilia dortmanna L. Limnology and Oceanography. 42(3): 529-537

Rodriguez, I. 2000. Ecohydrology: A hydrologic perspective of climate-soil-vegetation dynamics. Water

resources research. 36(1): 3-9

Rohde, H. & Rood, M.J. 1984. An evaluation of sulfate in European precipitation 1955-1982. Atmospheric

environment. 18: 2627-2639

Rudd, J. W. M., Kelly, C. A. and Schindler, D. W. 1988. Comment on “Dynamic model of in-lake alkalinity

generation” by L. A. Baker and P.L Brezonik. Water resources research. 24(10): 1825-1827

Paper 4

19

Rudd, J. W. M., Kelly, C. A. St. Louis, V., Hesslein, R. H., Furutani, A. and Holoka, M. H. 1986. Microbial

consumption of nitric and sulfuric acids in acidified north temperate lakes1. Limnology and

Oceanography. 31(6): 1267-1280

Sand-Jensen, K. and Prahl, C. 1982. Oxygen-exchange with the lacunae and across leaves and roots of the

submerged vascular macrophyte, Lobelia dortmanna L. New Phytologist. 91: 103-120

Scharling, M. 1999. Klimagrid – Danmark, Nedbør, lufttemperatur og potential fordampning 20*20 og 40*40

km. Danish Meteorological Institute. Technical report. 99-12

Schindler, D. W. 1988. Effects of acid rain on freshwater ecosystems. Science. 239: 149-157

Schindler, D. W., Turner, M. A., Staiton, M. P. and Linsey, G. A. 1986. Natural sources of acid neutralizing

capacity in low alkalinity lakes of the Precambrian Shield. Science. 232: 844-847

Schindler, D. W., Wagemann, R., Cook, R. B., Ruszzczynski, T., Prokopowich, J. 1980. Experimental

acidification of Lake 223, experimental lakes area: Background data and the first three years of

acidification. Canadian journal of fisheries and aquatic sciences. 37(3): 342-354

Seitzinger, S. P. 1988. Denitrification in freshwater and coastal marine ecosystems: Ecological and

geochemical significance. Limnology and Oceanography, 22(4/3): 702-724

Sigg, L., Johnson, A. and Kuhn, A. 1991. Redox conditions and alkalinity generation in a seasonally anoxic

lake (Lake Greifen). Marine Chemistry. 36: 9-26

Sutcliffe, D. W., Carrick, T. R., Heron, J., Rigg, E., Talling, J. F., Woof, C., Lund, J. W G. 1982. Long-term

and seasonal changes in the chemical composition of precipitation and surface waters of lakes and

tarns in the English Lake District. Freshwater Biology. 12(5): 451-506

Søndergaard, M. 1983. Heterotrophic utilization and decomposition of exstracellular carbon released by the

aquatic angiosperm Littorella uniflora (L.) ascers. Aquatic botany. 16(1): 59-73

Vestergaard. O. and Sand-Jensen, K. 2000. Aquatic macrophyte richness in Danish lakes in relation to

alkalinity, transparency, and lake area. Canadian Journal of Fisheries and Aquatic Sciences. 57: 2022-

2031

Williamson, C. E., Saros, J. E., Vincent, W. F., Smold, J. P. 2009. Lakes as reservoirs as sentinels,

integrators, and regulators of climate change. Limnology and Oceanography. 54(6/2): 2273-2282

Zalewski, M. 2002. Ecohydrology – the use of ecological and hydrological processes for sustainable

management of water resources. Hydrological Sciences. 47(5): 823 – 832

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