Dr. Kenneth A. Byrne Greenhouse Gas Emissions from Rewetted Peatland

Department of Life Sciences

Supervisor: Dr. Kenneth A. Byrne

Greenhouse gas emissions from rewetted peatland forests

Thesis presented by: Caitlin Rigney For the degree of DOCTOR OF PHILOSOPHY Submitted to the University of Limerick

November 2016

Abstract

Natural peatlands are important sinks of carbon (C) and vital in the global C cycle. Despite covering just 3% of the earth’s land mass, they store as much C as all terrestrial biomass. Drainage for forestry alters the hydrology and chemical reactions in peatlands, converting them from sinks to sources of carbon dioxide (CO2) and nitrous oxide (N2O), while reducing methane (CH4) emissions. Rewetting is considered an important tool in climate change mitigation and is utilized in addition to other management tools such as Sphagnum introduction to return the C sink function of peatlands and re- establish peat forming conditions in degraded peatlands. The first aim of this study was to investigate the controls on CO2, CH4 and N2O dynamics in two rewetted former peatland forest sites in Ireland; one blanket peatland eight years after rewetting (Pollagoona) and one raised peatland three years after rewetting (Scohaboy), produce annual greenhouse gas (GHG) balances for both peatlands and compare them with natural and forested systems. The second aim was to compare the chemical and physical properties of natural, drained and rewetted peatlands in order to assess the effect of both drainage and subsequent rewetting on peatland properties. Gas fluxes were measured using the chamber method. Micro sites comprising the dominant vegetation at the study site were established and gas balances produced for one year. Although interannual variablility influences GHG emissions from peatlands, results from this study represent fluxes in a year without extreme climatic episodes. In this study, both sites acted as CO2 and CH4 sources. Although Pollagoona was an overall CO2-C source -2 -1 - (131.6 g CO2-C m yr ), one microsite acted as a strong C sink (-142.84 g CO2-C m 2 -1 -2 -1 yr ). Methane emissions were small, totalling 2.94 ± 1.03 g CH4-C m yr . -2 -1 Molinia caerulea plots were both the greatest CO2-C (168.4 g CO2-C m yr ) and -2 -1 CH4-C (2.53±1.01 g CH4-C m yr ) sources. Pollagoona experienced N2O uptake (- 11.78 µg m-2yr-1) due to the behaviour of one microsite during the study period -2 -1 Scohaboy acted as a large CO2-C (585.3 g CO2-C m yr ) source in all microsites despite revegetation of non- brash plots due to the availability of fresh organic matter across the site. Scohaboy was also a CH4 source emitting 3.25 ± 0.58 g CH4-C -2 -1 -2 -1 -2 m yr . Both CO2-C (819.31 g CO2-C m yr ) and CH4-C (4.76 ±0.98 g CH4-C m

ii yr-1) emissions were highest from the brash plots. Annual nitrous oxide losses were small over the study period (72 µg m-2yr-1). Pollagoona displays little variability in elemental composition between land use type, while significant differences were observed in C and N content between land uses in Scohaboy. Unexpectedly N content of the peat did not increase with depth in four of the six peat cores analysed and this is reflected in their C:N ratio. All cores contained both labile and recalcitrant OM. Refractory OM was found in forestry samples from both sites and Scohaboy natural. Comparisons between land use types indicate that drainage and subsequent rewetting alter the properties analysed in this study.

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Declaration I ,the undersigned, declare that the work in this project is entirely my own and to the best of my knowledge contains no material previously written, published or submitted for merit or award by this university or any other academic establishment, except where acknowledgement has been made in the text.

Signed: Date:

Acknowledgments

I am grateful to my supervisor Dr. Kenneth Byrne for his help, time and guidance throughout this study.

Many thanks to Dr. David Wilson and Dr. Flo Renou Wilson for their time, experience, assistance, advice, patience and hospitality. Staying with you and having your knowledge to call upon was invaluable during this project.

I am indebted to my ‘minions’ (Jonay, DJ, Mike, Roisin, Sinead, Eva and Marilyne) who were called upon to assist with field work on a regular basis. Jonay, thank you for your help particularly while I was writing my thesis, it is much appreciated.

Thank you to the Environmental Protection Agency who funded the study by PhD fellowship (2012-CCRP- PhD. 2).

Special thanks to Ger who encouraged and supported me during my PhD, no matter what hour of the day I rang!

Thanks to my fellow postgrads both in the Schrodinger and Foundation building for help and friendship throughout the study.

I am grateful to my family for their support and encouragement during the last three years despite not understanding most of what I was talking about!

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Table of contents

Contents Abstract ...... i Declaration ...... v Acknowledgments ...... vii Table of contents ...... ix List of figures ...... xv List of tables ...... xvi Chapter One ...... 1 Introduction...... 1 Chapter Two ...... 7 Literature Review ...... 7 2.1 Global Overview of Peatlands ...... 9 2.1.1 Land Use of Peatlands ...... 10 2.1.2 Peatlands in Ireland ...... 12 2.1.3 Types of Peatland in Ireland ...... 13 2.2 Carbon and Greenhouse Gas Balance in Natural Peatlands ...... 15 2.2.1 Carbon gas dynamics in peatlands ...... 16 2.2.2 Carbon Dioxide ...... 17 2.2.3 Methane ...... 20 2.2.4 Nitrous oxide ...... 23 2.2.5 Waterborne carbon...... 24 2.3 Effect of drainage and changed land use on peatlands ...... 26 2.3.1 Agriculture,...... 27 2.3.2 Forestry ...... 29 2.3.3 Peat Extraction ...... 34 2.4 Rewetting and Restoration ...... 36 2.4.1 Rewetting and Restoration activities internationally ...... 38 2.4.2 Rewetting and Restoration activities in Ireland to date ...... 42 2.4.3 Impact of rewetting and restoration on peatlands...... 45 2.4.4 Methods of determining success or failure of rewetting and restoration ...... 49 2.5 Impacts of drainage and land use change and rewetting on peat properties ...... 50 2.5.1 Bulk Density ...... 50 2.5.2 Ash content ...... 51 2.6.1 International policy on peatlands ...... 52 Chapter Three...... 55

Carbon dioxide flux dynamics in a blanket peatland forest eight years after rewetting ...... 55 3.1 Abstract ...... 57 3.2 Introduction ...... 58 3.3 Materials and Methods ...... 59 3.3.1 Site Description ...... 59 3.3.2 Study Site ...... 62

3.3.3 CO2 flux measurements ...... 64 3.3.4 Environmental Variables ...... 65 3.3.5 Green Area Index ...... 66 3.3.6 Modelling the annual development of leaf area index ...... 67 3.3.8 Data Analysis ...... 67

3.3.9 Modelling of CO2 flux ...... 67 3.8.4 Statistical and uncertainty analysis ...... 70

3.8.2 Reconstruction of CO2 fluxes ...... 71 3.4 Results ...... 71 3.4.1 Environmental variables ...... 71 3.4.2 Vegetation dynamics ...... 74

3.4.3 On-site CO2 fluxes ...... 75 3.4.4 Model performance ...... 79

3.4.5 Annual CO2 balance ...... 81 3.5 Discussion ...... 83

3.5.1 Controls of CO2 fluxes ...... 83

3.5.2 Temporal variation in CO2 exchange ...... 85 3.5.3 Carbon balance of Pollagoona ...... 86 3.6 Conclusions ...... 89 Acknowledgements ...... 89 Chapter Four ...... 91 Carbon dioxide emissions in a raised bog after clearfelling and rewetting ...... 91 4.1 Abstract ...... 93 4.2 Introduction ...... 95 4.3 Materials and Methods ...... 97 4.3.1 Site Description ...... 97 4.3.2 Study Site ...... 100

4.3.3 CO2 flux measurements ...... 101 4.3.4 Environmental Variables ...... 102 4.3.6 Green Area Index ...... 103 4.3.7 Modelling the annual development of leaf area index ...... 104 4.3.8 Data Analysis ...... 104

4.3.9 Reconstruction of CO2 fluxes ...... 104

4.3.10 Modelling of CO2 Exchange ...... 105 4.3.11 Statistical and uncertainty analysis ...... 107 4.4 Result ...... 108 4.4.1 Environmental Variables ...... 108

4.4.2 Measured CO2 fluxes ...... 110 4.4.3 Model performance ...... 114

4.4.4 Annual CO2 balance ...... 116 4.5 Discussion ...... 118

4.5.1 Controls of CO2 fluxes ...... 118

4.5.2 Temporal variation in CO2 exchange ...... 120 4.5.3 Carbon balance of Scohaboy Bog ...... 120

4.5.4 Effect of brash on CO2 emissions ...... 122 4.6 Conclusions ...... 123 Acknowledgements ...... 123 Chapter Five ...... 125 Methane and Nitrous Oxide flux dynamics in a blanket peatland forest eight years after rewetting ...... 125 5.1 Abstract ...... 127 5.2 Introduction ...... 129 5.3 Materials and Methods ...... 131 5.3.1 Site Description ...... 131 5.3.2 Study Site ...... 131

5.3.3 Measuring CH4 and N2O fluxes ...... 131 5.3.4 Environmental Variables ...... 133 5.3.5 Flux estimation and statistical analysis ...... 133

5.3.6 CH4 and N2O flux modelling ...... 135 5.4 Results ...... 135 5.4.1 Environmental variables ...... 135 5.4.2 Measured gas fluxes ...... 138 5.4.3 Calculated Gas fluxes ...... 141 5.5 Discussion ...... 143 5.6 Conclusions ...... 147 Acknowledgements ...... 147 Chapter Six ...... 149 Methane and Nitrous Oxide emissions in a raised bog after clearfelling and rewetting..... 149 6.2 Introduction ...... 153 6.3 Materials and Methods ...... 154

6.3.1 Site Description ...... 154 6.3.2 Study Site ...... 154

6.3.3 Measuring CH4 and N2O fluxes ...... 154 6.3.4 Environmental Variables ...... 156 6.3.5 Flux estimation and statistical analysis ...... 157

6.3.6 CH4 and N2O flux modelling ...... 158 6.4 Results ...... 159 6.4.1 Environmental variables ...... 159 6.4.2 Measured methane fluxes ...... 162 6.4.3 Calculated Gas fluxes ...... 165 6.5 Discussion ...... 167 6.6 Conclusions ...... 170 Acknowledgements ...... 171 Chapter Seven ...... 173 The impact of afforestation and subsequent rewetting on peat properties in blanket and raised bogs ...... 173 7.1 Abstract ...... 175 7.2 Introduction ...... 177 7.3 Materials and Methods ...... 179 7.3.1 Study Site ...... 179 7.3.1 Determining pH ...... 182 7.3.2 Bulk Density ...... 182 7.3.3 Elemental Analysis ...... 183 7.3.4 Thermal Analysis ...... 185 7.3.5 Ash Content ...... 186 7.4 Results ...... 187 7.4.1 Physical and Elemental Analysis of Natural, Forested and Rewetted Blanket and Raised Peat ...... 187 7.4.2 Thermal Properties of Natural, Forested and Rewetted Blanket Peat ...... 192 7.4.3 Thermal Properties of Natural, Forested and Rewetted Raised Peat ...... 200 7.5 Discussion ...... 208 7.5.1 Physical and Elemental Soil Properties ...... 208 8.5.2 Thermal Soil Properties ...... 211 7.6 Conclusion ...... 213 Acknowledgements ...... 213 Chapter Eight ...... 215 General Discussion ...... 215 8.1 Greenhouse Gas balances ...... 217

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8.2 Assessment of rewetting success ...... 220 8.3 Future Work ...... 222 References ...... 223 Appendix 1 ...... 283

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List of figures Figure 2.1: Schematic representation and classical difference between 'bog' and fen...... 15 Figure 2.2 Carbon cycle in undrained peatlands and those drained for forestry (Minkkinen et al., 2008)...... 30 Figure 3.1 Location of the Pollagoona bog site ...... 60 Figure 3.2 Site map of Pollagoona rewetted area (outlined in red) containing both forested and undrained peatland...... 65 Figure 3.3 Pollagoona microsites in May 2014, prior to summertime Molinia growth…….63 Figure 3.4 CO2 flux chamber and raising collar, equipped with a cooling system and fan. ..65 Figure 3.5 Climate data for Pollagoona, Co Clare, during the sample period...... 73 Figure 3.6 Interpolated water table depths in all study plots from March 2014 until February 2015...... 74 Figure 3.7 The seasonal development of Green area index in each of the 8 sample plots in Pollagoona...... 75 -2 -1 Figure 3.8 Observed Net Ecosystem Exchange (NEE) (mg CO2 m h ) and ecosystem -2 -1 respiration (RTOT) (mg CO2 m h ) within the Eriophorum vaginatum – Sphagnum and Cladonia portentosa - Calluna vulgaris communities at Pollagoona Co Clare...... 77 -2 -1 Figure 3.9 Observed Net Ecosystem Exchange (NEE) (mg CO2 m h ) and ecosystem -2 -1 respiration (RTOT) (mg CO2 m h ) within the Molinia caerulea dominated communities at Pollagoona, Co Clare. Fluxes were graphed on Julian Day...... 78 Figure 3.10 Measured RTOT in response to both temperature at 5cm depth and water table depth...... 79 Figure 3.11 Measured PG in response to PPFD, Green Area Index and temperature at 5cm depth...... 80 Figure 3.12 Relationship between observed and modelled RTOT and PG on Pollagoona ...... 80 Figure 3.13 Average monthly modelled gross photosynthesis (PG), ecosystem respiration -2 -2 (RTOT) and net ecosystem exchange (NEE) (g CO2- C m month ) for (a) Eriophorum- Sphagnum (b) Cladonia- Calluna, and (c) Molinia caerulea sample plots...... 82 Figure 4.1 Location of the Scohaboy bog site ...... 97 Figure 4.2 Site map of Scohaboy rewetted area (outlined in red)…………………………..98 Figure 4.3 Scohaboy microsites in May 2014……………………………………………100 Figure 4.4 CO2 flux chamber and raising collar, equipped with a cooling system and fan102 Figure 4.5 Climate data for Scohaboy Bog, Co Tipperary during the sample period, March 2014-March 2016...... 109 Figure 4.6 Mean sample day water table in brash plots, Sphagnum/ Eriophorum plots and Cladonia/ Mosses plots, March 2014- February 2015...... 110 -2 -1 Figure 4.7 Observed net ecosystem exchange (NEE) (mg CO2 m h ) and ecosystem -2 -1 respiration (RTOT) (mg CO2 m h ) within the Eriophorum vaginatum – Sphagnum and Cladonia portentosa- Mosses communities at Scohaboy Bog, Co. Tipperary...... 112 -2 -1 Figure 4.8 Observed net ecosystem exchange (NEE) (mg CO2 m h ) and ecosystem -2 -1 respiration (RTOT) (mg CO2 m h ) within the Eriophorum vaginatum and Brash communities at Scohaboy Bog, Co. Tipperary...... 113 Figure 4.9 Measured RTOT response to both water table depth and temperature at 5cm depth...... 114

Figure 4.10 Measured PG in response to water table depth and PPFD...... 115 Figure 4.11 Relationship between observed and modelled RTOT and PG on Scohaboy Bog 115 Figure 4.12 Average monthly modelled gross photosynthesis (PG), ecosystem respiration -2 -2 (RTOT) and net ecosystem exchange (NEE) (g CO2- C m month ) for (a)Eriophorum- Sphagnum (b) Cladonia- Mosses, (c) Eriophorum and (d) brash sample plots. s...... 117 Figure 5.1 CH4 and N2O equipment set up………………………………………...... 132 Figure 5.2 Climate data for Pollagoona, Co Clare, during the sample period...... 136 Figure 5.3 Interpolated water table depths in all study plots ...... 137

-2 -1 Figure 5.4 Observed CH4 fluxes (mg CH4 m h ) within the (a) Eriophorum vaginatum – Sphagnum (b) Cladonia portentosa - Calluna vulgaris and (c) Molinia communities at Pollagoona, Co. Clare ……………………………………………………………………138 -2 -1 Figure 5.5 Measured N2O fluxes (μg m hr ) within the (a) Eriophorum vaginatum – Sphagnum (b) Cladonia portentosa - Calluna vulgaris and (c) Molinia communities at Pollagoona, Co. Clare………………………………………………….…………………..140 -2 -1 Figure 5.6 Interpolated monthly means of CH4-C (g m month ) within the Eriophorum vaginatum – Sphagnum, Cladonia portentosa - Calluna vulgaris and Molinia communities at Pollagoona, Co. Clare…………………………………...... 141 Figure 6.1 CH4 and N2O equipment set up……………………………………………….156 Figure 6.2 Climate data for Scoghaboy Bog Co. Tipperary……………………… …160 Figure 6.3 Mean sample day water table in brash plots………………………...... 161 Figure 6.4 Observed CH4 fluxes (mg CH4 m-2 h-1) within the (a) Eriophorum vaginatum- Sphagnum, (b) Cladonia portentosa- Mosses and (c) Brash communities at Scohaboy Bog, Co.Tipperary……………………………………………………………………………….162 Figure 6.5 Measured N2O fluxes (μg m-2hr-1) within the (a) Eriophorum vaginatum- Sphagnum, (b) Cladonia portentosa- Mosses and (c) Brash communities at Scohaboy Bog, Co. Tipperary…….164 Figure 6.6 Interpolated monthly means of CH4-C (g m-2 month-1) within the Eriophorum vaginatum- Sphagnum, Cladonia portentosa- Mosses and Brash communities at Scohaboy Bog, Co. Tipperary………………………………………………………………………...165 Figure 7.1 DTG curves of (a) natural, (b) forested and (c) rewetted blanket peatland at Pollagoona………………………………………………………………………………….192 Figure 7.2 DSC curves of (a) natural, (b) forested and (c) rewetted blanket peatland at Pollagoona………….………………………………………………………………………193 Figure 7.3 Evolution of percentages of Exo1, Exo2 and Exo2 fractions with respect to total OM in (a) natural,(b) forested and (c) rewetted blanket peatland at Pollagoona………………………………………………………………………………….194 Figure 7.4 Recalcitrance levels in (a) natural (b) forested and (c) rewetted blanket peatland at Pollagoona…………………………………………………………………………………196 Figure 7.5 Heat of combustion (QSOM) levels in (a) natural (b) forested and (c) rewetted blanket peatland at Pollagoona…………………………………………………………….197 Figure 7.6 Thermal stability in (a) natural (b) forested and (c) rewetted blanket peatland at Pollagoona………………………………………………………………………………….198 Figure 7.7 DTG curves of (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy…………………………………………………………………………………...200 Figure 7.8 DSC curves of (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy…………………………………………………………………………………...202 Figure 7.9 Evolution of percentages of Exo1, Exo2 and Exo3 fractions with respect to total OM in (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy…………………………………………………………………………………...203 Figure 7.10 Recalcitrance levels in (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy…………………………………………………………………………………204 Figure 7.11 Heat of combustion (QSOM) levels in (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy……………………………………………………………….205 Figure 7.12 Thermal stability in (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy…………………………………………………………………………………207

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List of tables

-2 -1 Table 2.1 Greenhouse gas fluxes (CO2-C CH4-C N2O-N g m yr ) in peatlands used for agriculture ...... 28 -2 -1 Table 2.2 IPCC emission factors (CO2-C CH4-C N2O-N kg ha yr ) for peatlands used for forestry. Positive values indicate a flux from peat to the atmosphere...... 31 -1 -1 Table 2.3 Greenhouse gas fluxes (CO2-C, CH4-C, N2O-N tonnes ha yr ) from peat extraction areas...... 35 Table 2.4 Peatland Restoration Guidance documents ...... 41 -1 -1 Table 2.5 Carbon gas fluxes (CO2-C, CH4-C, g m yr ) from rewetted peat extraction, agricultural and forestry sites...... 47 Table 2.6 Published bulk density values from natural peatland and drained peatland forestry ...... 51 Table 3.1 Vegetation species recorded in the study plots. Species are listed in descending order of dominance...... 63 2 Table 3.2 Estimated parameter values and goodness of fit (r ) for PG ...... 69 2 Table 3.3 Estimated parameter values and goodness of fit (r ) for RTOT ...... 70 Table 3.4 Summary of annual sums or averages of variables from the Eriophorum- Sphagnum (b) Cladonia- Calluna, and (c) Molinia caerulea sample plots...... 83 Table 4.1 Vegetation species recorded in the study plots. Species are listed in descending order of dominance……………………………………………………………………...... 99 2 Table 4. 2 Estimated parameter values, goodness of fit (r ) for PG...... 106 2 Table 4.3 Estimated parameter values, goodness of fit (r ), for RTOT ...... 107 Table 4.4 Summary of annual sums or averages of variables from the sample plots. Eriophorum- Sphagnum, Cladonia- Mosses, Eriophorum and brash sample plots...... 118 Table 5.1 Vegetation species recorded in the study plots. Species are listed in descending order of dominance...... 134 Table 5.2 Summary of annual sums or averages of CH4 and N2O and relevant variables from the three study microsites………………………………………………………………………………………………..142 Table 6.1 Vegetation species recorded in the study plots. Species are listed in descending order of dominance...... 158 Table 6.2 Summary of annual sums or averages of CH4 and N2O and relevant variables from the three study microsites…………………………………………………………………………………….166

Table 7.1 Summary of soil bulk density (g dm3) for Pollagoona and Scohaboy in each treatment ...... 187 Table 7.2 Ash content (%) per treatment in Pollagoona and Scohaboy in 10 cm increments ...... 188 Table 7.3 Summary of soil pH for Pollagoona and Scohaboy in each treatment...... 189 Table 7.4 Summary of soil carbon content (%) for Pollagoona and Scohaboy in each treatment...... 189 Table 7.5 Summary of soil nitrogen content (%) for Pollagoona and Scohaboy in each treatment...... 190 Table 7.6 Summary of C: N ratio for Pollagoona and Scohaboy in each treatment...... 190 Table 7.7 Summary of soil sulphur content (%) for Pollagoona and Scohaboy in each treatment ...... 190

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Chapter One

Introduction

Although peatlands cover just 3% of the world’s land surface they store as much carbon (C) as all terrestrial biomass and double that of global forest biomass (Parish et al., 2008). Carbon stocks in peatland are estimated to be over 400 Pg (Kaat & Joosten, 2008; Yu, 2011), the majority of which is found in the northern hemisphere (Strack, 2008a). In peatlands, decomposition is impeded by saturated anaerobic conditions caused by a high water table leading to accumulation of organic matter (Moore, 1987; Renou-Wilson et al., 2011). The C cycle in peatlands involves carbon dioxide (CO2) up take by plants in photosynthesis, plant and soil respiration where

CO2 is released and methane (CH4) production and consumption. Peatlands provide many services such as supporting biodiversity, regulating water flow, C sink (Peatland Ecology Research Group, 2009), fuel source, providing land for agriculture (Joosten and Clarke, 2002), archaeology and amenity (Kniess and Trepel 2007; Moors for the Future Partnership, 2012). In the undrained condition, peatlands act as sinks of CO2 (Laine et al., 2006; Aurela et al., 2009), having a net cooling impact on the atmosphere. Drainage precedes peatland utilization for agriculture, forestry and fuel extraction. Lowering the water table alters microbial activity, nutrient conditions, increases aeration of the surface peat thereby increasing decomposition and C loss, causes subsidence and increases peat bulk density (Minkkinen and Laine et al., 1998; Laine et al., 2006). Greenhouse gas (GHG) dynamics are affected by the falling water table as CO2 emissions increase, CH4 emissions decrease and N2O is produced (Martikainen et al., 1995; Silvolia et al., 1996; Drewer et al., 2010; Turetsky, et al., 2014), converting them from a sink to a source of both C and N2O, causing them to have a net warming impact on the atmosphere. Rewetting and restoration are seen as viable means of restoring the C sink function of peatlands (Höper et al., 2008). The primary aim of peatland restoration must be to limit further peat degradation (Schuman and Joosten, 2008) and subsequently re- establish an ecosystem similar to the one which was degraded (Konvalinková and Prach, 2014). In order to accomplish either of these, drainage ditches must be blocked to raise the water table (Bragg, 2011) and if successfully rewetted, recolonization of the area by peatland species will occur, eventually leading to C accumulation (Komulainen et al., 1998). Three options exist for restoration practitioners (a) to allow the peatland to recover spontaneously, (b) to direct the spontaneous recovery and succession towards a desired target and (c) to implement

3 specific restoration measures (Konvalinková and Prach, 2014). Rewetting often leads to a decrease in both CO2 and N2O emissions and an increase in CH4 emissions (Tuittila et al., 2000b; Waddington and Price, 2000: Regina and Myllys, 2009; Beyer and Höper, 2015), although these changes can take years to occur, as sites may persist as C sources after rewetting (Petrone et al., 2003; Samartitani et al., 2011). Peatland restoration is a relatively new area of research that has increased in popularity since the 1990s (Rochefort et al., 2003) with an increased awareness of the ecosystem services a peatland provides. Although previous studies across the globe have contributed to the present body of knowledge surrounding peatland restoration (Sliva & Pfadenhauer, 1999; Tuittila et al., 1999; Wilcox et al., 2006; Poschlod et al., 2007; Armstrong et al., 2009; Whinam et al., 2010; Mahmood & Strack 2011; Xiaohong et al., 2012), few have dealt with any aspect of rewetting peatland forests (e.g. Komulainen et al., 1999; Haapalehto et al., 2011; Koskinen et al., 2016). In Ireland, significant work on GHG dynamics of both rewetted peat extraction (Wilson et al., 2007b, Wilson et al., 2009; Wilson et al., 2016b) and agricultural peatland (Renou- Wilson et al., 2016) has informed estimates of emissions from temperate rewetted sites but GHG fluxes in rewetted forests have not been studied. Estimates of emissions from rewetted peatland forests are vital to develop nationally specific C emission factors for inclusion in the national GHG inventory. For this study two rewetted forest sites were selected, one blanket and one raised bog. Previously published literature values will be used for comparative purposes. Eight plots were identified per site representative of the different vegetation communities and topography on each site. Using chamber methods CO2, CH4 and nitrous oxide (N2O) fluxes were measured at regular intervals over an 18 month time period (March 2014 – September 2015). Peat properties in both study sites and in adjacent natural and forested sites were also analysed. The overall aim of the study was to measure the biosphere-atmosphere exchange of greenhouse gases CO2, CH4 and N2O and to investigate the relationship of these gases with environmental and climatic variables such as soil temperature, water table, photosynthetic photon flux density and vegetation. The measurements were made at the vegetation community level as this is the level of spatial variation in GHG gas exchange. The study was divided into five complementary sub- projects with each one having specific aims:

1. Carbon dioxide flux dynamics in a blanket peatland forest 10 years after rewetting (Chapter 3)

(a) To identify the controlling variables of CO2 fluxes on rewetted blanket peatland forest sites,

(b) To estimate the annual CO2 balances for a range of rewetted microsites at a rewetted blanket peatland forest ten years after rewetting (c) To compare the findings with other rewetted peatland sites and suggest

management options for increasing the CO2 sink potential.

2. Carbon dioxide emissions in a raised bog after clearfelling and rewetting (Chapter 4)

(a) Identify the controlling variables of CO2 fluxes on such rewetted raised peatland forest sites,

(b) Estimate the annual CO2 balances for a range of rewetted microsites at a newly rewetted raised peatland forest in the temperate region

suggest management options for increasing the CO2 sink potential.

3. Methane and nitrous oxide emissions in a blanket peatland forest 10 years after rewetting (Chapter 5)

(a) Measure CH4 and N2O dynamics in rewetted blanket peatland

(b) Determine had emissions returned to those of natural peatlands

4. Methane and nitrous oxide flux dynamics in a raised bog after clearfelling and rewetting (Chapter 6)

(a) Measure CH4 and N2O dynamics in rewetted raised peatland

(b) Determine had emissions returned to those of natural peatlands

5. The impact of afforestation and subsequent rewetting on peat properties in blanket and raised bogs (Chapter 7)

(a) Compare the bulk density, ash content, pH and elemental composition of natural, forested and rewetted blanket and raised peatland in Ireland.

(b) Conduct thermal analysis (DTG and DSC) in order to assess the variety in thermal stability and quality of OM between the three treatments sites

Chapter Two

Literature Review

2.1 Global Overview of Peatlands

Peatlands are wetland ecosystems where decomposition is impeded under saturated, anaerobic conditions leading to the accumulation of organic matter, deriving from decaying plant and animal material, as peat (Moore, 1987; Renou-Wilson et al., 2011). Lack of oxygen in the peat disrupts the enzymatic activity required for efficient decomposition (Freeman et al., 2001). Consequently peatlands store large amounts of C and nitrogen (N) in their natural state. Carbon is stored because more C is produced in photosynthesis than released in respiration, i.e. biomass production rate exceeds the rate of decomposition in an anaerobic environment (Clarke and Joosten, 2002; Byrne, et al., 2004). Interaction between plant productivity and C loses through decay, leaching and deposition of C into the mineral soil below the peat layer contributes to the rate of peat accumulation (Clarke and Joosten, 2002). Optimal peat accumulation occurs in a humid climate (Mäukiläu, 1997). Natural peatlands are characterised by a high water table, anaerobic soil environment, partial decomposition of the deposited organic material on the surface and are located in areas with moderate (750-1,000 mm per annum) to high rainfall levels (>1,200 mm) (Otte, 2003; Renou and Farrell, 2005).

Peatlands occupy a key position in regulating the global C cycle through both their effect on the terrestrial C store and influence on atmospheric greenhouse gas concentrations. Although a large C store in their natural condition, peatlands are susceptible to climatic and hydrological changes which can switch their role from sink to source of C to the atmosphere (Strack, 2008a). Estimates of the global peat C store vary, ranging from 357-882 Pg (Bolin et al., 1979; Eswaran et al., 1993; Woodwell and Mackenzie et al 1995) although more recent estimates of the peatland C stock have a smaller range 450 (Kaat & Joosten, 2008) - 612 Pg of carbon (Yu, 2011), most of which occurs in the northern hemisphere (Strack, 2008a). Estimating C stocks in peat remains difficult due to the absence of comparable, detailed data on peatland location, uncertainties in the depth of peat on site and bulk density values used, underestimation of the world’s largest peatlands and of shallow peats (Turunen et al. 2002; Kuhry and Turunen 2006; Yu, 2011; Yu, 2012). Peatlands regularly occurring in a complicated mosaic in the landscape together with other soil types,

9 can be difficult to identify from an aerial view and may be unreachable from the ground. Peatlands act to cool the climate by reducing CO2 concentration in the atmosphere (Frolking et al., 2006); however they are a significant natural CH4 source (Crill et al., 1992; Laine et al., 2007b). All three greenhouse gases emitted from peatlands, CO2, CH4 and N2O have individual greenhouse warming potentials (GWP). Global warming potential is used to compare the total energy that a gas will absorb over a set period of time. Carbon dioxide functions as the baseline value for GWP with a value of 1 (EPA, Unites States Environmental Protection Agency,

2013). Nitrous Oxide has a GWP 265–298 times that of CO2 for a 100-year timescale (EPA, Unites States Environmental Protection Agency, 2013). Methane has a global warming potential about 65 times higher than CO2 over 20 years

(Ramaswamy et al., 2001) or 21 fold higher that CO2 (Joabbson et al., 1999; Forster et al., 2007) on a 100 year time scale. Wetlands, to which peatlands belong, globally emit between 80 and 280 Tg (Bartlett and Harris, 1993; Mathews, 1996; Bridgham et al., 2013), supplying 20-30% of the global CH4 emissions (Houghton et al., 2001).

2.1.1 Land Use of Peatlands

Peatlands are estimated to cover approximately 4 million km2 of the earth’s surface (Joosten, 2009). Up to 90% of the world's peatlands are located in the boreal, temperate and subarctic zones of the northern hemisphere, while the remainder is found in tropical areas (Feehan and O’Donovan, 1996). According to Montanarella et al. (2006), peatlands in Europe total in the region of 329,000 km2, as estimated using the European Soil Database; almost half of which occur in the north western region of the continent. Central and eastern European countries (excluding Russia) account for almost 60 000 km2 of European peatlands (Montanarella et al., 2006; Minayeva et al., 2009). Many of these peatlands are forested or paludified land, differing from those peatlands found in other parts of Europe (Bragg and Lindsey 2003; Minayeva et al., 2009). Bogs found in oceanic regions of the world such as north-western Europe, eastern North America and Alaska are not forested, supporting a mix of Sphagnum mosses and evergreen shrubs (Moore, 2002). They often consist of raised domes or open plateaux (Moore, 2002; Renou et al., 2006). Raised bogs elsewhere are commonly forested (Seppä, 1996).

Peatlands provide many services such as supporting biodiversity, regulating water flow, acting as a carbon sink (Peatland Ecology Research Group, 2009), source of fuel and land for agriculture (Joosten and Clarke, 2002). Other functions include archaeology and amenity (Kniess and Trepel 2007; Moors for the Future Partnership, 2012). Some of these services provided by peatlands (i.e. fuel source, agriculture and forestry) when utilized, lead to anthropogenic interference and a reduction of the capability of the peatland to function as a carbon sink, regulate water flow or maintain local biodiversity. Once disturbed, peatlands will emit significant volumes of the greenhouse gas, CO2 and in some cases N2O (Martikainen et al., 1995; Byrne, et al., 2004). It has been found that all countries contaiing peatlands are net GHG emitters from those peatlands. The levels of emissions due to the land use of organic soils in individual countries are uncertain (Byrne, et al., 2004). Estimates of global annual CO2 emissions from drained peatland are 1.3 Gt (excluding emissions from peat fires) (Joosten, 2010).

Companies and agencies which use the resources supplied by peatlands are now turning to rewetting and restoration as a means of reducing the environmental impact of anthropogenic use of peatlands. Through their actions they endeavour to return the peat accumulating ability to damaged peatlands (Quinty and Rochefort 2003). Before reviewing the restorative work completed on Irish peatlands to date, it is first necessary to consider the terminology associated with this work and what it means in practise. Restoration is used to encompass almost all rehabilitative and restorative work on Irish peatlands. Care must be taken so as not to confuse the language; rewetting, rehabilitation and restoration. Rewetting a peatland involves human interventions, i.e. drain blocking, or allowing natural development which leads to a rise in the water table. Restoring or restoration suggests returning a peatland ecosystem to a former state of being or re-establishing the ecosystem that was there prior to damage and degradation (Renou-Wilson, 2011). However, restoration can also indicate the aim is to return the peatland to a point along it’s developmental course which was interrupted when drainage and development began (Vasander et al., 2003). Rehabilitation, as used by some agencies conducting restorative work on peatland, may in fact be more accurate to describe the results of projects carried out on many peatlands to date. Rehabilitation describes peatlands following rewetting and restoration work which act as a carbon sink but do not demonstrate species

11 composition, community structure or the peatland ecosystem present immediately prior to drainage, plantation, extraction etc. (Bord na Móna, 2010a). It may demonstrate a possible state that was present on the peatland at some point in its past development, however, it is not known if these sites will ever return to fully functioning peatlands. For the purpose of this review, the term rewetting will be used to describe the activities conducted on the study sites as the work conducted centred on re-establishing the water table on the sites.

2.1.2 Peatlands in Ireland

Irish peatlands feature prominently in the Irish landscape and have been estimated to cover between 17% (Hammond, 1981) and 20% (Connolly and Holden, 2009) of the total land area. They account for 53% (Tomlinson, 2005) to 61% (Xu et al., 2011) of the total soil C stocks in the country. Therefore they have the potential to play a vital role in the future carbon budgeting of this country. Initially thought of as a means to provide for future generations through reclamation and agriculture, peatlands have become recognised as a global carbon sink, i.e. peatlands store carbon (Gorham, 1991). While Irish peatland has been estimated to currently total 1,205,235 ha (Eaton, et al., 2008), total values vary as reclamation and land use have influenced assessments. The primary uses of peatland in Ireland consist of peat extraction, agriculture, forestry, wind farming, recreation and tourism (Renou-Wilson, et al., 2011). Total acreage of some land uses are conflicting with authors reporting agriculture on peatland to range from 89,600 ha (modified from Oleszczuk, et al., 2008) to 295,000 ha (NPWS, 2013), and that under forestry stretching from 222,998 ha (Malone and O’ Connell, 2009) to 300,000 ha (Black, et al., 2008).

Three types of peatland are prevalent in Ireland, raised bog found in the Midlands, blanket bog occurring in western and mountainous regions (Renou and Farrell, 2005). and fen peat. Each one owes its differences to climate, drainage, water source, geology, nutrient input and man’s activities (Hammond, 1981).

Ireland is situated between the 51° and 55° northern latitudes and as such experiences a moderate, temperate climate of cool wet summers and mild winters.

Therefore climatic conditions are ideal for peat formation with up to one fifth of the country being covered by peat soils (Renou- Wilson and Byrne, 2015). The significance of Irish peatlands cannot be overlooked as they cover almost 1.5 million ha or 20 % of the land of Ireland (Connolly and Holden, 2009). Peatlands are wetland ecosystems which are characterised by the accumulation of organic matter deriving from decaying plant and animal material under saturated, anaerobic conditions (Renou-Wilson et al., 2011). Decomposition of the organic matter is slowed down and impeded in the anaerobic environment present due to the high water tables found on peatlands (Moore, 1987).

2.1.3 Types of Peatland in Ireland

The Midlands of Ireland is a flat, low lying plain underlain by limestone rock, complete with deposits of glacial drift. This has influenced greatly the drainage and nutrient status of the ground waters of the area, thereby manipulating the conditions during the initial stages of peat formation (Hammond, 1981). Raised bogs are found throughout this low lying calcareous plain which occupies the midland area (Otte, 2003). Raised bogs are domed in shape, elevated above their nearby landscape and receive their water input solely from atmospheric depositions thereby causing them to be nutrient poor (Fig 2.1). These peatlands are thereby labelled as ombrotrophic (Gorham, 1991; Laine and Vasander, 1996; Päivänen and Hänell, 2012). Irish raised bogs are unique and differ from those occurring elsewhere due to their treeless condition and their less domed shape (Renou et al., 2006). Raised bogs elsewhere are commonly forested (Seppä, 1996). Plant species commonly found on raised bog includes Sphagnum, Calluna (IPCC n.d.)

Blanket bog is found primarily along the Western seaboard and at high elevations where precipitation levels are high (Black et al., 2008). It too is ombrotrophic peat which receives nutrients and water from both atmospheric deposition and precipitation (Laine et al., 2006). Blanket peat development is controlled by a climate of cool summer, high rainfall (greater than 1,250mm and no less than 225 rain days per annum) and high humidity (Hammond, 1981). The history of blanket bogs started 10,000 years ago. Originally, peat formation was confined to shallow lakes and hollows in the landscape. In time, acid peat spread from these areas,

13 forming a blanket covering the surrounding landscape. Heavy rainfall following the change in climate 4500 years ago led to leaching of the surface soil and the formation of an iron pan lower down in the soil. As a result, the whole area became water logged and peat formation continued (Foss and O’Connell, 1998).

There are two types of blanket bog, Atlantic; confined to coastal plains and mountain lowlands no higher than 200 m above sea level, and Mountain blanket bog; found at higher elevations on slopes with inclines at less than 15° (Foss et al., 2001; Otte, 2003). Common species found on blanket peat include Sphagnum sp, Calluna vulgaris, Molinia caerulea, Tricophorum, Eriophorum sp, Schoenus nigricans, Empetrum nigrum and Vaccinium myrtillus (IPCC n.d.; Conaghan, 2000, Black et al., 2008).

A fen is also peatland which differs in nutrient levels and acidity to blanket and raised bog types. With greater nutrients available and lower acidity to other bog types, fen peat is able to support a greater plant and animal community (Foss, 2007). It occurs where there is a permanently high water table. A fens’ primary source of water is groundwater or surface water, causing the elevated calcium and nutrient levels and fens can also be labelled minerotrophic peatland (Laine and Vasander, 1996; Irish Peatland Conservation Council, 2002; Renou-Wilson et al., 2011). Fens and bogs share common ancestry (Fig 2.1). After the end of the last Ice Age, about 10,000 years ago, many shallow lakes were formed and became vegetated over time. This vegetation produced litter which fell to the floor of the lake. In the oxygen deficient environment at the bottom of the lake, decomposition was impeded. A built up of partially decomposed organic litter occurred, filling the lake basin. A constant water supply to the lake basin from the surrounding mineral rich ground ensured the continued formation of the peat. Eventually the peat extended out of the original lake basin, cutting off the supply of mineral rich water to the basin. At this stage of its development, the accumulated peat is classified as fen peat. Further accumulation of peat continued on some fens under the influence of nutrient poor rainwater, leading to ombrotrophic bog formation (Laine and Vasander, 1996; Otte, 2003).

Following the fen stage, plants invaded the nutrient poor habitats. Sphagnum moss or bog moss was one of the most established peatland invaders where it acted as a sponge drawing water up to the surface of the bog keeping it water logged. About

4500 years ago, annual rainfall decreased and with the subsequent drying of the bog surfaces came the establishment of pine woodland which remained until the climate changed about 500 years ago and annual rainfall increased, leading to rapid bog growth in the following years (Foss and O’Connell, 1998).

Figure 2.1: Schematic representation and classical difference between 'bog' and fen. Brown = peat; arrow = water flow (adapted from Renou-Wilson et al., 2011).

2.2 Carbon and Greenhouse Gas Balance in Natural Peatlands

Peatlands contribute significantly to the world C cycle (Yu et al., 2011). Due to their structure and formation, peatlands contain a huge reservoir of carbon (Tomlinson 2005; Strack 2008a; Gorham, 1991). Many authors have shown that natural peatland areas act as a sink for CO2 (Shurpali et al., 1995; Maljanen et al., 2010; Koehler et al., 2011; Helfter at al., 2015). Studies conducted in Sweden found a mean net CO2 −2 −1 exchange of −78 g CO2 m yr (Lund et al., 2007) (CO2 uptake by the peatland is −2 −1 indicated by the negative value) for nutrient poor sites and −189 g CO2 m yr (Nilsson et al., 2008), for minerotrophic sites. Minerotrophic sites in northern −2 Finland returned fluxes varying between −188 and −219 g CO2 m (Aurela et al., 2007). Despite acting as a C sink, emission levels from unmanaged sites can fluctuate greatly depending on weather conditions and location (Aurela et al., 2004; Maljanen et al., 2010).

For many years, anthropogenic activities have contributed to increases in GHG’s in the atmosphere (Khalil 1999). Anthropogenic causes of GHG increases in the atmosphere take many forms, include disturbing natural peatlands, thereby converting them from a sink to a source of CO2 (Nykänen et al., 1995). Agriculture, forestry and peat extraction for fuel and horticultural use are major causes of peatland disturbance in Ireland (Renou- Wilson et al., 2011). These activities involve extensive drainage and ploughing which trigger significant changes in the functionality and sink function of peatlands (Kasimir- Klemedtsson et al., 1997), impacting on emissions (Renou- Wilson et al., 2015; Wilson et al., 2015; Renou- Wilson et al., 2016). The most prominent problems appearing after drainage of peatland are subsidence due to mineralization, increased soil compaction and loss of

CO2 (Flessa et al., 1998).

In general, it has been found that the greatest effect on the elemental components of the peat is the water table (Haapalehto, et al., 2011). The dynamics of peatlands are extremely sensitive to changes in the hydrological cycle (Erwin, 2009). Peat moisture content is one of the most important factors limiting decay of the organic material in peatlands (Clarke and Joosten, 2002). Diffusion of gases in water is limited due to the lower than ambient temperatures caused by waters’ large heat capacity. This contributes to the low availability of oxygen in the waterlogged peatland, reducing the rate of decay of organic material and causing peat accumulation (Clarke and Joosten, 2002).

2.2.1 Carbon gas dynamics in peatlands

The carbon cycle in peatlands starts with the assimilation of carbon by plants through photosynthesis. The predominant plants in most peatlands responsible for this are Sphagnum spp. shrubs and sedges. Plant respiration releases approximately

50% of the CO2 back to the atmosphere following photosynthesis while the remaining portion is stored as organic matter (Ryan, 1991). Litter is produced both above (leaves) and below (roots) the surface of the peat. The soil profile is divided into two distinct layers due to peatland hydrology. The acrotelm is the surface layer where water table fluctuates, rooting occurs and aerobic conditions are present

(Quinty and Rochefort, 2003). Above the water level, the aerobic conditions allow the rapid decomposition of litter and release of CO2 (Moore and Dalva, 1997). The catotelm lies under the acrotelm and is permanently waterlogged, thus microbial activity and decomposition of material remains slow due to the aerobic conditions present (Quinty and Rochefort, 2003). As the litter decomposes it’s density increases and structure breaks down (Päivänen and Hånell, 2012). As decomposition continues, mass is lost from the litter which is covered by fresh litter and over time the gradually rising water level envelopes it, adding the litter to the anaerobic catotelm (Clymo, 1984). In the acrotelm, C losses are due to CO2 emissions and much less to CH4. Due to the incomplete decay of the organic matter, a portion of it accumulates in the anaerobic catotelm. Anaerobic decomposition in the catotelm continues slowly and is conducted by methanogens which results in the production and emission of CH4 (Couwenberg, 2009). Methane produced in the catotelm is normally susceptible to oxidation in the acrotelm (Päivänen and Hånell, 2012). Carbon accumulation rates in a peatland vary temporally, as litter deposition and decomposition rates are dependent on the plants present and climatic conditions, both of which have varied throughout the last 10,000 years (Charman et al., 2015). As decomposition continues in the catotelm, C storage in the peat relies on the rate of C input to the catotelm being greater than catotelm losses.

2.2.2 Carbon Dioxide

Carbon dioxide exchange is affected simultaneously by several biotic and abiotic factors such as temperature, water table, plant species present, light levels, microtopography and restrictions on enzyme activation (Bubier et al., 1998; Christensen et al., 1998; McKenzie et al., 1998; Freeman et al., 2001; Tuittila et al., 2004). Net ecosystem exchange (NEE) in a peatland is composed of two major processes, respiration and photosynthesis. Temperature exercises great control over carbon dioxide emissions from soils (Moore and Knowles, 1989; Silvola et al., 1996; Helfter et al., 2015). Ecosystem respiration in peatland is strongly correlated to soil temperature (Lafluer et al., 2005) and more often than water table, it is the dominating control of variations in respiration (Lafleur et al. 2005; Mäkiranta et al.

2009; Juszczak et al., 2013). Microbial activity is stimulated at higher air and soil temperatures creating higher CO2 emissions (e.g. Frolking and Crill 1994; Silvola et al. 1996). Despite the strong correlation with temperature, the sensitivity of ecosystem respiration to it varies greatly within and between sites (Drösler et al. 2006; Mäkiranta et al. 2009). Heterotrophic respiration may respond to temperature differently depending on the chemical composition of nutrient availability and substrates (Updegraff et al. 2001; Blodau et al. 2004). Carbon dioxide emissions are frequently linked to temperature at a particular depth in the soil profile, often at 5 cm (Maljanen et al., 2003; Minkkinen et al., 2007). Temperature recordings at staggered depths down through the soil profile e.g. at 5 cm, 10 cm and 20 cm are necessary to connect the gas fluxes measured to the environmental conditions present on site (Tuittila et al., 2004). Although it works as an emissions regulator simultaneously with water table, following water level draw down, temperature has a reduced effect on respiration (Silvola et al., 1996).

The importance of water table in regulating soil respiration has not reached a consensus in the scientific literature. Many studies have supported the dominant role of water table in respiration regulation (e.g. Silvola et al. 1996; Bubier et al. 2003; Tuittila et al. 2004; Riutta et al. 2007), while others have rejected it (Lafleur et al. 2005; Dimitrov et al. 2010). Water table is seen by many as a primary regulator of gas exchange in peatlands (Tuittila et al., 1999; Glatzel et al., 2006; Wilson et al.,

2007). It acts as a control for CO2 emissions in conjunction with soil temperature (Moore and Dalva, 1993). It is responsible for determining the aerobic or anaerobic condition of the peat and the depth of the oxic zone. A high water table is required to create anoxic conditions, thereby causing low respiration rates and incomplete decomposition of the organic matter in peatlands (Moore 1989; Clarke and Joosten, 2002). The anaerobic conditions have been credited with controlling the release of C to the atmosphere by restricting the function of enzymes which would hasten the breakdown of organic matter (Freeman et al., 2001). As easily decomposable organic matter is stored in peat surface layers, even a short term lowering of the water table could lead to rapid C loss through oxidation (Alm et al., 2007). Ecosystem respiration follows the water table gradient of highest respiration when the water table is deepest (Laine et al., 2007a). Once the water table is lowered to a certain depth, soil respiration no longer increases in response (Silvolia et al., 1996;

Chimner and Cooper, 2003) and in extremely dry conditions peat respiration rate may decline, such as in a study in Finland where it was found that respiration declined when the water level fell below 61 cm (Mäkiranta et al., 2009).

Photosynthesis is controlled by such factors as photosynthetically active radiation, water table, vegetation, temperature and CO2 concentration (Townsend et al., 2003). The water table of a peatland regulates the amount and type of vegetation present.

Peatland vegetation binds atmospheric CO2 during photosynthesis while respiration and the consumption of soil organic matter by animals and microorganisms releases

CO2 back into the atmosphere (Alm et al., 1997). Altering the water table in a peatland causes changes in the vegetation on site. Net primary productivity particularly during the growing season is highly susceptible to hydrological changes (Sottocornola et al., 2010). Mosses in particular will react negatively to dry conditions (Dimitrov et al., 2011). Distinct vegetation communities will react differently to changes in the hydrologic regime. In drier conditions, photosynthesis in shrubs may increase while more water-reliant plants, e.g. mosses and sedge will experience reduced photosynthesis (Bubier et al 2003; Strack and Price 2009). Water table depth also affects vegetation succession in the peatland; Sphagnum species production increases with high water tables while deep water tables encourage shrub growth (Potvin et al., 2015). Over time, vegetation succession impacts photosynthesis and net primary production (NPP) (Strack et al., 2006a). Each vegetation type has an individual rate of photosynthesis; sedges have a much higher rate of photosynthesis than deciduous species while evergreen species exhibit the lowest rates of photosynthesis (Bubier et al., 1998; Moore et al., 2002b). When C uptake decreases at warmer temperatures, it is suspected to be caused by increased respiration or suppressed photosynthesis (Aurela et al., 2009). Considering all the above interactions a peatland may act as a carbon sink or source in climatically different years (Laine et al., 2009).

Net primary production, which indicates the rate of CO2 uptake, is chiefly regulated by PAR, peatland hydrology and the availability of nutrients, mainly N and phosphorus (P) (Blodau, 2002). Peat pH affects NPP by controlling or limiting the nutrient availability to plants (Limpens et al., 2008). Vegetation present on bogs reflects this with bog specialists displaying adaptations such as ever greenness,

19 thickening of the plant epidermis, high root biomass, recycling nutrients, forming mycorrhiza and having a long lifespan (Mitsch 2009; Rydin and Jeglum, 2013).

Carbon dioxide fluxes are measured in order to estimate the net ecosystem exchange (NEE) of the system. Carbon uptake and loss are estimated based on the fluxes measured and relationships determined between environmental factors and gas emissions (Alm et al., 1997). In temperate peatlands in the south west of Ireland, -2 Laine et al., (2006) determined an undrained site to be a sink of 206-242 g CO2 m -1 -2 -1 yr , in Scotland, a sink of 324-500 g CO2 m yr was determined by Drewer et al., (2010) and in Finland under varying summer conditions a peatland was deemed a -2 -1 sink of 12- 216 g CO2 m yr over a three year period (Aurela et al., 2009). Despite different conditions and weather patterns, all sites acted as sinks over the course of the study years.

2.2.3 Methane

Natural peatlands are a considerable source of CH4 (Sundh et al., 1994). Natural wetlands account for 80% of natural CH4 emissions and 1/3 of all global emissions (Ciais et al., 2013). Methane is released from peat through three primary methods; ebullition, diffusion or plant facilitated transport (Joabsson et al., 1999; Lai, 2009). Methane is produced by the decomposition processes found in peat (Martikainen et al., 1995). Anaerobic methanogenic bacteria produce CH4 in the anoxic layer of the peatland as the last product of organic matter decomposition. They primarily reduce

CO2 with molecular hydrogen and dismutate acetate (a root exudate), to CO2 and

CH4 (Whalen, 2005; Conrad, 2007; Nazaries et al., 2013). Aerobic methanotrophic bacteria present in the near surface oxidise CH4 (Whalen, 2005) consuming CH4 as a carbon and energy source (Murrell 2010). Therefore, the CH4 flux on the peatland is the balance achieved between these two combating processes (Sundh et al., 1994) and the type of CH4 transport to the atmosphere (Couwenberg, 2009). In addition to these microbial processes, CH4 fluxes on peatlands are connected with peat aeration, nutrient level, vegetation cover (particularly the presence of aerenchyma species), peat density, peat temperature, pH and water table level (Williams and Crawford, 1984; Dise, 1993; Nykänen et al., 1998; Ström et al., 2005; Armstrong et al., 2015).

Methane production levels remains challenging to predict due to the opposing controls on methanogensis (microbial production) vs methanotrophy (consumption) in peatland ecosystems (Turetsky et al., 2014).

Peatland vegetation plays a crucial role in CH4 emissions (Ström et al., 2003; Henneberger et al., 2015a; Henneberger et al., 2015b) by influencing the processes in

CH4 cycling (Joabsson et al., 1999). Plants have adapted to aquatic habitats by developing aerenchyma which allow O2 transport to roots in anoxic soil (Armstrong et al., 1991), thus prompting CH4 oxidation by methanotrophs in the peat rhizosphere

(Whalen, 2005). Aerenchyma also provides a direct route for the release of CH4 to the atmosphere, avoiding the methanogenic bacteria (Whalen 2005; Green and Baird,

2011). Different vegetation communities greatly influence CH4 emissions on a peatland. Beyer and Höper, (2014), found that rewetted sample sites dominated by

Eriophorum vaginatum and E. angustifolium displayed higher CH4 balances compared to Ericaceae, Molinia Juncus or Sphagnum dominated sites with a comparable water level. This is attributable to the leaves of E. vaginatum and E. angustifolium facilitating CH4 transport through the aerenchym (Joabbson et al., 1999; Joabbson and Christensen, 2001; Drösler, 2005; Couwenberg et al., 2011). In addition, decaying vegetation and root exudates increase the substrate available for heterotrophic microorganisms and eventually for methanogens (King and Reeburgh,

2002; Saarnio et al., 2004). Species specific differences in NEE and CH4 emissions have been observed for common peatland sedges, Carex, Eriophorum and Juncus explained by variations in root exudation arrangement and radial oxygen loss (Ding et al., 2005: Ström et al., 2005; Koelbener et al., 2009). Variations in the active methanotrophic and methanogenic bacteria communities connected with Carex and Eriophorum spp. have been established (Cheema et al. 2015). Mosses also likely influence methanotrophic activity and lead to CH4 oxidation (Liebner at al., 2011; Franchini et al., 2015).

Water table impacts significantly on the volume of CH4 produced from peatland. It is one of the major factors controlling the rate of both CH4 production and oxidation (Komulainen et al., 1998; Joabbson and Christensen, 2001). The area of highest methanogenic activity is normally located close to the water table, at the boundary between the aerobic and anaerobic layers (Segers, 1998). Therefore, in situations with a high water table, methanogenic bacteria are active near the peat surface,

21 reducing transport distance to the surface and opportunity for oxidation. Methane emissions decrease as the water table decreases in a peatland due to the presence of oxygen which restrains the action of the CH4 producers (Nykänen et al., 1998). Emissions from peatlands where water table levels are below -20cm have been found to be insignificant, however, once water tables rise, emissions increase (Huttunen et al., 2003b; Couwenberg et al., 2011; Beetz et al., 2013). This reduction of CH4 can be attributed to the larger zone of oxidation between the water table and the surface, a longer pathway to the atmosphere for CH4 allowing more oxidation, less readily available substrate in the lower peat layers and the inaccessibility of vegetation mediated transport as the water table lies below the rooting zone of aerenchymous plants (≈ 20-30cm depth) such as Eriophrium (Roulet et al., 1993; Nykänen et al., 1998; Alm et al., 1998; Strack et al., 2006b; Audet et al., 2013). Water levels which are above the peat surface also contribute to low CH4 emissions (Couwenberg et al.,

2011). Flooded conditions most likely constrain CH4 production by inhibiting CH4 diffusion through stationery water, reduced vascular plant biomass contributing to a decrease in CH4 production and where water is moving, a probable rise in O2 availability (Turetsky et al., 2014).

Methane production is also temperature dependent (Dunfield et al., 1993) and therefore temperature is an important regulator of seasonal changes in CH4 emissions (Huttunen et al., 2003a). Temperature increases in peat cause greater volumes of

CH4 to be produced (Dunfield et al., 1993; Moore and Dalva 1993). As soil temperature increases, so too does the methanotrophic (consumption) and methanogenic (production) activity. At temperatures above 15°C, methanotrophs can no longer compensate for the high production of CH4 allowing more to be transported to the surface (van Winden et al., 2012). Furthermore, temperature may increase CH4 transport by stimulating promoting ebullition through an rise in bubble volume and distribution of CH4to gaseous phase (Fechner-Levy and Hemond, 1996), in addition to encouraging plant-mediated transport through increased pressurized ventilation and subsequently CH4 ﬂow in the aerenchyma (Große,1996) thereby leading to higher rates of CH4 emissions. The pH of a peatland influences GHG emissions, as acidity can possibly restrain methanotrophic and methanogenic action. Many methanogenic bacteria will grow in the narrow range of pH 6-8 (Garcia et al., 2000). Certain acidophilic methanogens have been found to grow at pH 5.5 (Kamal

22 and Varma, 2008). The optimal pH for CH4 production has been found to be near neutral, between 5 and 8.5 (Bubier et al., 1993; Kamal and Varma, 2008). These optimum conditions are rarely found in the field and the low pH limits CH4 production as demonstrated by Dunfield et al., (1993) when CH4 emissions were increased by raising the pH of peat.

2.2.4 Nitrous oxide

The predominant natural source of nitrous oxide (N2O) is soils. Although peatlands contain approximately 30% of the world’s soils organic N reserve fluxes of N2O are very low or insignificant in natural peatland (Martikainen et al., 1993; Regina et al., 1996; Minkkinen et al., 2002). Nitrous oxide is produced during both nitrification (aerobic) and denitrification (anaerobic) processes by bacteria in the soil (Davidson and Schimel, 1995). Denitrification can be categorised as a microbiologically ‘broad process’ which can be conducted by a wide variety of microbes in comparison to the ‘narrow process’ of autotrophic nitrification (Butterbach-Bahl et al., 2013). Denitrification is an anaerobic process, principally involving the bacterial genera PseudoMónas, Micrococcus, Bacillus, and Thiobacillus in which C functions as the source of energy and nitrogen the electron acceptor (Kamal and Varma, 2008).

During the denitrification process, nitrogen is reduced to the gases N2O and N2 (Zumft, 1993; Ye et al., 1994). Nitrous oxide may also be produced during + nitrification through oxidation of NH4 (Koops et al., 1997).

Given that N2O production is driven by microbial reactions, soil temperature is vital in controlling the N2O emission rate (Lohila et al., 2010). Nitrous oxide emissions generally increase with temperature (Szukics et al., 2010), although this relationship has been found to be inconsistent and at times non-significant (Lohila et al., 2010;

Juszczak and Augustin 2013). Water table also contributes to the governance of N2O emissions in peatlands. Peatlands with high water tables release small amounts of

N2O (Drewer et al., 2010) due to the anoxic conditions. The optimal moisture condition for N2O emissions is 55% water filled pore space while at higher saturation levels N2O emissions are inhibited (Szukics et al., 2010). Overall, draining peatlands, thereby lowering the water table, causes an increase in N2O emissions

(Regina et al., 1996; Huttunen et al., 2003a), although this is dependent on their nutrient status (Regina et al., 1996).

Nitrous oxide production is correlated to the C:N ratio in the soil C and N pools (Ambus et al., 2006). Increased use of N fertilizers over recent decades may have contributed significantly to increased N2O concentrations in the atmosphere (Khalil,

1999). Augustin et al., (1998) found that fertilizer use contributed to heightened N2O emissions from peat. Nitrous oxide emissions have also been found to be influenced by both the oxygen and nutrient status of the site (Martikainen et al., 1993).

2.2.5 Waterborne carbon

Besides gaseous losses to the atmosphere, C is also lost through waterborne carbon fluxes such as dissolved organic carbon (DOC), particulate organic carbon (POC), dissolved inorganic carbon (DIC) and gaseous CO2 and CH4 , in run- off from the peatland (Evans et al., 2015). Surface waters from peatland catchments, supersaturated with CO2 and CH4 represent a significant pathway connecting C store with the atmosphere (Dinsmore et al., 2009a). Dissolved organic carbon is the largest component of water borne carbon from a peatland (Olefeldt et al., 2013). It is described as organic carbon that is smaller than 0.45 µm in diameter (Thurman, 1985). Significant geophysical and ecological roles are filled by DOC in both peatland and downstream environments affecting nutrient availability, acidity, metal mobility and light penetration in aquatic ecosystems (Steinberg, 2003 ).Waterborne carbon can contribute greatly to the carbon balance of a peatland, turning it from a calculated sink or neutral peatland to a carbon source (Roulet et al., 2007). Dissolved organic carbon, similarly to CO2 originates in the aerobic area above the water table. Aerobic decomposition generates compounds easily transported when the water table rises following rainfall. The regulators of DOC production are not fully understood but involve soil chemistry, soil temperature and the activity of microbes (Holden, 2005; Koehler et al., 2009). A significant relationship has been determined between vegetation type and DOC concentration levels (Armstrong et al., 2012), where Calluna was associated with the highest DOC, sedges with intermediate DOC and Sphagnum spp with the lowest DOC. This association may be due to the behaviour of the water table in the areas these plants colonise. Lowering the water table of a

24 peatland causes a pulse of DOC losses from the system (Strack et al., 2008b). Once DOC enters the drainage system, it is vulnerable to photochemical breakdown (Köhler, et al., 2002; Cory et al., 2014) and may be exploited as a C and energy source by microorganisms (Battin et al., 2008). In a peatland in south west Ireland, the annual export of DOC was 14.1 (±1.5) g C m−2 for the study year (Koehler et al.,

2009). Significant relationships have been found between organic carbon and CO2 emissions in lake water in Ireland (Whitfield et al., 2011).

Particulate organic carbon is defined as organic matter greater than 0.7µm (Hope et al., 1997; Dawson et al., 2004). Particulate organic carbon exports are normally small in natural and managed peatlands, its production is primarily controlled by vegetation cover (Worrell and Evans, 2009; Evans et al., 2015). Where bare peat surfaces are exposed, POC fluxes account for the largest proportion of fluvial C losses (Pawson et al., 2008) up 100g C m-2 year -1, as found by Worrell et al., (2011), from an eroding blanket bog. Drainage alone, without additional exposed peat surfaces, has not been found to create large POC fluxes (Evans et al., 2015), however erosion of the ditches or gullies themselves may be a source of large POC losses (Evans et al., 2006).

Dissolved CO2 and DIC in water routinely originate from soil respiration and mineral weathering (Worrell et al., 2005). Dissolved inorganic carbon, that is C from − 2 − carbonate sources, (H2CO3 + HCO3 + CO3 ) (Worrall et al., 2003), is important in aquatic systems as it buffers against swift changes of pH and regulates the amount of inorganic carbon available for photosynthesis, it can precipitate out of the water as

CaCO3, removing inorganic carbon from the water (Kalff, 2002). Dissolved inorganic carbon levels are connected to the peatland’s age; as a peatland ages, levels decrease (Löfgren 2011). Dissolved inorganic carbon fluxes in waters flowing from bogs are generally low due to the low solubility of CO2 at low pH (Evans et al., - 2015). However, in fen peats, where higher pH is present and linked to HCO3 and 2- CO3 , DIC fluxes are higher (e.g. Worrell et al., 2003; Fiedler et al., 2008). Weather events have been shown to influence waterborne C, the levels of gaseous CO2 and

CH4 in particular increasing after rainfall events (Waddington and Roulet, 1997).

2.3 Effect of drainage and changed land use on peatlands

Drainage precedes most land use changes on peatland. Lowering of the water table alters fundamental ecosystem characteristics such as microbial activity, aeration of the surface peat and nutrient conditions (Laine et al., 2006). The hydrological functioning of the peatlands is threatened during the reclamation process. Drainage will potentially increase decomposition in the soil and loss of C. Bulk density often increases as subsidence occurs (Minkkinen and Laine et al., 1998). The most prominent problems appearing after drainage of peatland are subsidence due to mineralization, increased soil compaction and loss of CO2 (Flessa et al., 1998). Subsidence is loss of soil height, as a result of the physical process the peatland undergoes and the oxidation of the peat. Initially after drainage, settlement of the layers above the water table may occur followed by the consolidation of the lower layers (Richardson et al., 1991). The rate of subsidence varies with peat type, rate of decomposition, thickness of the peat layer, climate of the area and land use (Oleszczuk et al., 2008). Shrinkage of the peat will occur after this stage. Lowering the water table also leads to a loss in buoyancy in the upper soil horizons, causing compression of the peat layers below the water table. Therefore drainage of the soil results in the sinking of the surface from a combination of shrinking and compaction (Kasimir- Klemedtsson et al., 1997; Oleszczuk et al., 2008). As the soil continues to dry out, vertical and horizontal shrinkage cracks will appear, allowing the diffusion of oxygen into the deeper layers of the soil profile (Oleszczuk et al., 2008). This increases aerobic respiration and peat decomposition at these levels. The depth of the oxic zone in the peatland is increased by lowering the water table, encouraging the growth of aerobic microbial decomposers (Chmielewski, 1991), which encourages aerobic respiration, increasing the rate of aerobic decomposition and thereby increases CO2 losses to the atmosphere (Wilson, 1996; Rydin and Jeglum 2006; Jaatinen et al., 2007; Preston et al., 2012).

The falling water table increases the availability of oxygen in the peat, causing the production of N2O, rather than N2 (Drewer et al., 2010). Fen peatland

(minerotrophic) shows an increased N2O flux following drainage whereas ombrotrophic sites have shown little change before and after drainage likely due to their nutrient poor nature. High nitrogen supply increases the rate of N2O fluxes

(Nykänen 1995; Regina 1996; Regina et al., 1998; Nykänen et al., 2002). Nitrate oxidizers have a higher presence on drained peatland than their natural counterparts (Regina, 1996). Drainage and therefore lowering the water table by 1cm has been -2 -1 found to increase CO2 emissions by 7.1 mg CO2 m yr at certain temperatures in comparison to undrained sites (Silvolia et al., 1996). Methane emissions respond to the falling water levels. Levels of CH4 produced on drained peatlands are greatly reduced when compared to undrained peatlands (Martikainen et al., 1995; Turetsky, et al., 2014). Soil management, will contribute to the extent of shrinking and compaction (Nykänen et al., 1995). As with all peatland, reclaimed land has the ability to store large amounts of water and minerals are plentiful to plants in the initial years after drainage, due to peat mineralisation. Reclaimed and drained peatlands are more likely to emit increased CO2 but less CH4 than natural peatlands (Moore and Knowles 1989; Silvola et al., 1996; Von Arnold et al., 2005). Following drainage, greater volumes of DOC are lost than from undrained peatlands (Strack et al., 2008).

2.3.1 Agriculture,

As stated above, peatland must be drained before many land use changes, as aeration and water conditions are changed to meet the requirements of agriculture. Grassland management systems and drainage of peatland accelerate the oxidation of the formerly accumulated organic matter, stimulating CO2 and N2O emissions while reducing methanogensis (Byrne et al., 2004; Freibauer et al., 2004; Maljanen et al., 2004; Nieveen et al., 2005; Jacobs et al., 2007; Elsgaard et al., 2012; Schrier-Uijl et al., 2013). Agricultural intensity, occurrence of ploughing and fertilization will contribute to the extent of shrinking and compaction which occurs on the peatland (Nykänen et al., 1995).

-2 -1 Table 2.1 Greenhouse gas fluxes (CO2-C CH4-C N2O-N g m yr ) in peatlands used for agriculture. Positive values indicate a flux from peat to the atmosphere.

CO2-C CH4-C N2O-N Land Use Reference source Location g m-2 yr-1 g m-2 yr-1 g m-2 yr-1 Finland 830 -0.1 0.848 Barley Maljanen et al., 2004 Finland 568 -0.052 1.108 Barley Maljanen et al., 2007

Finland 340 - - Barley Lohila, 2008 Finland 395 -0.1 0.275 Grassland Maljanen et al., 2004 Netherlands 220±90 - - Grassland Jacobs et al., 2007

Ireland 233 0 0.16 Grassland Renou- Wilson et al., 2014 Finland 405 0.071 0.574 Grassland Maljanen et al., 2007

The nutrient and drainage status of a peatland will impact on NEE, biomass and fluvial C loss, while local climate and management practices also influence the annual C balance (Maljanen et al., 2004; Renou- Wilson et al., 2014). Lower than average agricultural peatland CO2 emissions have been attributed to management practices; extensive grazing, no fertilization and water tables of -25cm (Renou- Wilson et al., 2016), confirming that agricultural emissions can be maintained at relatively low levels by maintaining a mean annual water table of -25cm (Regina et al., 2015). Soil cultivation and drainage are the most damaging of agricultural practices, which trigger huge shifts in emissions of gases from reclaimed land (Kasimir-Klemedtsson et al., 1997). Peatlands sown with rotational arable crops lose more C than those in permanent grassland (Elsgaard et al., 2012). Carbon dioxide emissions from arable peatland soils can be as much as 20% higher than peatland used for grassland and pastures (Maljanen et al., 2007). In arable peatlands, N2O emissions increase following rainfall and ploughing events (Maljanen et al., 2002; Maljanen at al., 2004; Eldar and Lal, 2008). Nutrient rich agricultural peatlands exhibit higher C losses (gas emissions and fluvial C) than their nutrient poor counterparts (Renou-Wilson et al., 2014; Barry et al., 2016). Nutrient rich peatland, -1 -1 converted to arable land can produce up to 41,100 kg CO2 ha a , with high variation expected (Oleszczuk et al., 2008). Frequent tillage of the soil keeps the top

28 layers of the soil oxygenated thereby enhancing decomposition of the peat and maintaining high CO2 emission rates (Nykänen et al., 1996). Reducing tillage or conversion to no till methods can have positive effects on CO2 and N2O emissions from agricultural peatland (Eldar and Lal, 2008).

Agricultural activity on reclaimed land in Ireland has mainly consisted of grassland production and animal grazing. Grassland covers approximately 4.3 m ha of Ireland (CSO, 2012), of that organic soils total about 300 000 ha. In particular, blanket bogs have been critically overgrazed (Renou-Wilson et al., 2011). In addition to increased peat decomposition, peatland reclamation for agriculture results in the almost complete elimination of the peatland flora and fauna. Organisms which thrive in the peatland environment die off when water table is lowered and are replaced with different communities, often with fewer plant species (Irish Peatland Conservation Council, 2009). High animal stocking rates contribute to the impact on C dynamics associated with agriculture in three key ways; removal of vegetation cover reduces the amount of photosynthesis taking place, thereby reducing the carbon sequestering capacity of the peatland, organic matter input available for peat formation is reduced and livestock on the land will cause further compaction of the peat (Garnett et al., 2000).

2.3.2 Forestry

Some peatlands support tree growth in their natural state however the high water table stunts root growth and peatland forest stands are unevenly sized and aged (Päivänen and Hånell, 2012). Therefore drainage is carried out to increase productivity in treed peatlands and create more hospitable growing conditions in treeless peatlands. The water table is lowered by inserting a series of ditches or drains. Drainage is a vital element of preparation for forestry in as the water table on the majority of peatland sites is too high for successful tree growth. The roots of terrestrial plant species cannot live in the normally waterlogged soil (Alm, et al., 2011).

Figure 2.2 Carbon cycle in undrained peatlands and those drained for forestry (Minkkinen et al., 2008).

Following forest drainage the C cycle and processes in the peatland are altered. As outlined earlier, the water table plays a fundamental role in regulating all aspects of gas exchange. With the reduction of the water table, changes to CO2, CH4 and N2O exchange are initiated. While the CO2 soil emissions increase, as the forest grows and develops, tree biomass increases, contributing to soil C through improved litter production, underground biomass production and the contribution of recolonising vegetation (Laiho and Finer 1996; Domisch et al., 1998; Hargreaves et al., 2003; Laiho et al., 2003). However, in concurrence with this, increased aeration of the peat causes an increase in organic matter decomposition rate, leading to increased CO2 emissions from the soil (Minkkinen et al., 2008). Estimates of C loss over a first

-1 forestry rotaton range from 1 to 14.6 t CO2 per ha per year (Hargreaves et al. 2003;

Byrne and Farrell 2006; Minkkinen 1999; Alm et al. 2007). Methane emissions are reduced and peatlands may even be converted to a CH4 sink (Roulet and Moore,

1995; Alm et al., 2011). N2O emissions generally increase following forestry drainage, depending on the trophy of the peatland (Martikainen et al., 1995; Aerts and Ludwig, 1997; Ojanen et al., 2010). Time since drainage can have an effect on soil nutrient pools, particularly magnesium and iron (Westman and Laiho et al., 2003). Following drainage, intensified oxidation of the organic material releases hydrogen ions into the peatland, increasing acidity (Laine et al., 1995), causing pH to decrease, particularly on nutrient rich sites (Westman and Laiho, 2003).

-2 -1 Table 2.2 IPCC emission factors (CO2-C, CH4-C, N2O-N kg ha yr ) for peatlands used for forestry. Positive values indicate a flux from peat to the atmosphere.

a b Location CO2-C CH4-C N2O-N Boreal kg ha-2 yr-1 kg ha-2 yr-1 kg ha-2 yr-1 Nutrient Poor 250 5.25 0.22 Nutrient Rich 930 1.5 3.2 Temperate 2600 1.88 2.8 -2 -1 a Values have been converted to kg ha yr , b Values have been converted to CH4- C. (IPCC, 2014)

Draining a peatland for forestry often changes the conditions where vegetation is growing dramatically (Laine and Vanha- Majamaa, 1992). Typical bog plants such as Eriophorum vaginatum, Eriophorum angustifolium, Calluna vulgaris, Sphagnum species, Vaccinium oxycoccus and Molinia caerulea to name but a few are specially adapted to the water logged nutrient poor conditions. When their habitat is drained, these plants will disappear to be replaced by woodland species (Korpela 1999; Laine and Vanha- Majamaa 1992). Carbon dioxide fluxes in peatlands are affected by the seasonal development of the plants present (Alm et al., 1997; Bubier et al., 1998; Tuittila et al., 1999). This vegetation, present on the peatland, influences soil respiration rate and photosynthesis at the site (Alm et al., 1997). Peatland vegetation binds atmospheric CO2 during photosynthesis, after which some is released through respiration or contributes to the formation of new plant tissue. A quantity of C is stored as CO2 by litter input below the ground while a proportion is released through heterotrophic respiration and the consumption of soil organic matter by animals and microorganisms (Alm et al., 1997; Wilson et al., 2007b,).

The impact of forestry on peatlands varies in line with the intensity of forestry management and soil disturbance on site (Groot, 1998; Pietilainen and Moilanen, 2003; Minkkinen et al., 2008). Some harvesting may occur on undrained or barely drained sites while other forestry sites have had a greater impact on the hydrology and carbon dynamic of the peatland through drainage, planting and fertilizing (Schumann and Joosten, 2008). Clear-felling of forested peatland raises the water table and raises the temperature of the peat, two processes known to affect greenhouse gas emissions in peatland. In the short term, clear-felling causes a rise in nitrous oxide emissions (Huttunen et al., 2003a). Fertilization is often required for tree growth on peatlands. Forestry establishment on blanket peat is difficult at times as it is very acidic and has low fertility. Nutrients essential for growth, in particular, P and N are in short supply with small volumes of N reserves in the soil (Renou and Farrell, 2005). Fertilization with N and P is vital to ensure tree viability (Minkkinen et al., 2008), otherwise growth rate will decline (Byrne and Farrell, 1998). Fertilization has been proven to cause increases in nitrous oxide emissions from peat (Augustin et al., 1998).

The first attempts at reclaiming European peatlands for forestry started in the 18th century (Zehetmayer 1954). Approximately 15 m ha of peatland worldwide is estimated to be drained for forestry, the majority of that in Fennoscandia and the former USSR (Paavilainen and Päivänen, 1995). Peatlands have a long association with forestry as they are seen as sources for renewable resources of wood (Päivänen and Hånell, 2012).

Some 730,000 ha or 10% of Irish land is forested (Forest Service 2012), of which approximately 322, 000 ha is on peatland (NFI, 2013), planted with species such as Sitka spruce (Picea sitchensis [Bong] Carr), lodgepole pine (Pinus contorta Dougl.), Norway spruce (Picea abies (L.) (Karst) and Scot’s pine (Pinus sylvestris L.), (Renou- Wilson, 2011; Renou-Wilson and Byrne 2015) at a time when peatland was seen as marginal and of little use. The 1960s, 1970s and 1980s saw a period of government driven reforestation in Ireland, very often on marginal, ‘waste’ peatland. Throughout the west of the country this occurred on blanket peat, culminating in more than 200 000 ha of blanket peats being afforested (NFI, 2013). Throughout the midlands, cutover and raised bog was afforested with up to 82 920 ha (Black et al., 2008) of conifer species. Preparation of peatlands in Ireland for forestry in the past

32 typically included ploughing and has been replaced by mounding, involving the use of soil from drains to form elevated mounds that are used for planting (Minkkinen, et al., 2008). Afforestation of peatlands is now limited, emphasis is on reforestation or other management options. Forestry establishment on blanket peat in particular is difficult at times; blanket bog is very acidic and has low fertility. Initially seen as means of utilizing marginal land and providing local employment, public and government support for forestry on peatlands has declined in recent years (Coillte, 2008 a; Black et al., 2009). Economically, the costs of maintaining viable forestry on peatlands are high and a growing public awareness of the environmental impacts of drainage, fertilization and clearfelling as well as the value of peatlands in their natural state has contributed to this (Holden et al, 2004; Renou and Farrell, 2005).

While plantation forests in Ireland are a significant store of carbon and have the potential to sequester considerable amounts of C in the future in the growing trees (Byrne and Milne, 2006), this must be balanced against questions of peatland forestry’s economic viability, lack of knowledge of C dynamics beyond the first rotation and the obligation to protect designated habitats and species under EU law (Wilson et al., 2013). The net carbon sink of a peatland is reduced by afforestation (Byrne and Milne, 2006), due predominantly to the higher oxidation and degradation levels in the peat caused by the lowering of the water table. Peatlands are increasingly valued as a unique habitat and natural resource as well contributing significantly to the Irish national carbon budget due to their sink function (NPWS 2013). Therefore, in some instances rewetting and restoration of unproductive peatland may be more pertinent than reforestation. Despite the Forestry Act of 1946 which requires replanting of forestry areas following clearfelling (Government of Ireland, n.d.), permission has been granted to Coillte not to replant some areas with restoration potential (Coillte, n.d.) following clearfelling.

2.3.3 Peat Extraction

In recent decades, peat extraction has been carried out either by small scale processes (tractor pulled hopper machine) or large scale industrial processes such as milled peat procedures (Wilson, et al., 2012). Peat extraction at any intensity, fundamentally impacts the ecosystem functioning of a peatland. In industrial scale peat extraction drainage ditches are installed at 15 m intervals to allow for the use of heavy machinery. Lowering the water table of the peatland causes an increase in peat oxidation and a subsequent rise in CO2 emissions (Holmgren et al., 2006). Removing surface vegetation and upper layers of the peatland eliminates photosynthesis and consequently the CO2 fixing capacity of the peatland (Nykänen et al., 1996; Waddington and Price, 2000), transforming the peatland from a CO2 sink and source of CH4 (natural condition), to a large CO2 source with diminished CH4 emissions (Sundh et al., 2000) (see table 2.3.). Carbon dioxide emissions are dependent on climate, time since drainage and the quality of the peat (Nykänen et al.,

1996; Alm et al., 1997; Waddington et al., 2002). Although CH4 emissions from the peat may diminish, emissions from the drainage ditches may still be substantial, particularly if they revegetate (Sundh et al., 2000; Hyvönen et al., 2013). In natural peatlands, N2O emissions are insignificant, however, following drainage for peat extraction emissions may increase considerably (Augustin et al., 1998), particularly in nutrient rich peatlands (Regina et al., 1996). Prominent problems appearing after drainage of peatland are subsidence due to mineralization and increased soil compaction (Flessa et al., 1998). As is well known in peatland research, water table depth impacts on the production of both CO2 and CH4.

Following cessation of peat extraction, peatland sites will continue to emit CO2 however these levels can be reduced with human intervention by raising the water table and careful management (Waddington et al., 2002). Simultaneously, N2O emissions may decrease while CH4 emissions increase (Augustin et al., 1998).

The peatlands of Ireland have been utilized as a source of energy since before the 8th century (Holmgren et al., 2008). World War II prompted rapid expansion of peat production, much of it hand cut, for use as fuel for the Irish population (RTE, 2015), and large scale peat extraction continued in the post war years (Feehan and

O’Donovan 1996). Bord na Móna are the primary peat producers in Ireland, owning up to 80% of the peatland used for extraction and harvesting in the region of 4 million tonnes per annum (Bord na Móna, 2010b). Between Bord na Móna and private peat producers, it is estimated that 100,000 ha of Irish peatlands are being utilized for peat production (Fitzgerald 2006). However, the quantity of peatlands affected by domestic peat production remains unknown (Renou-Wilson et al., 2011). Peat is utilized in the production of 8.5 % of Irish electricity (Forfas, 2010).

-1 -1 Table 2.3 Greenhouse gas fluxes (CO2-C, CH4-C, N2O-N tonnes ha yr ) from peat extraction areas. Positive values indicate a flux from peat to the atmosphere.

CO2-C CH4-C N2O-N Reference source Location Tonnes ha-1 Tonnes ha-1 Tonnes ha-1 yr-1 yr-1 yr-1 Finland 2.4 0.002 0.0002 Nykänen et al., 1996

Sweden 0.55-2.73 0.003- - Sundh et al., 2000 0.034a Canada 3.98b - - Waddington and Warner 2001 Canada 0.88-3.97b - - Waddington et al., 2002

Canada 3.02 0.014 - Cleary et al., 2005 Sweden 2.73c 0.02 0-0.016 Holmgren et al., 2006

Finland 3.16 0.004 0 Holmgren et al., 2006

Ireland 1.9-3.5 -0.001 - Wilson et al., 2007 Finland 1.89-11.18 0.054 0.002 Alm et al., 2007

IPCCd 0.2-1.1 0 0.001-0.002 Penman et al., 2003

a Includes emissions from drainage ditches, b May- August period only, c Includes emissions from stockpiles, d IPCC default factor for nutrient poor and nutrient rich industrial peatlands (CO2- C and

N2O- N) and for drained organic soils (CH4- C) (Wilson, et al., 2012).

Peat extraction has been conducted by milling since the 1970s (Renou et al., 2006). Large scale peat extraction on a site eventually creates one of two situations;

35 cutaway peatland where only 50cm -1m of peat remains over the underlying mineral soil or cutover peatlands where the remaining peat depths are greater (Rydin and Jeglum, 2013) and so after uses will vary for these sites. After uses of such cutaway sites will not include agricultural development as this had been deemed commercially unsustainable (Bord na Móna, 2010a), with other uses having to be found. Cutaway peatlands have been developed for commercial forestry with limitations following trials and have also been used to create woodland habitats and been allowed to naturally regenerate as wetlands, following some management (Wilson et al., 2013; Bord na Móna 2016).

2.4 Rewetting and Restoration

Restoration and rewetting is widely seen as a viable means of restoring the C sink function of peatlands (Höper et al., 2008). The aim of peatland restoration must first and foremost be to limit further peat degradation (Schuman and Joosten, 2008). Following this, the target remains the re-establishment of an ecosystem similar to the one which was degraded (Konvalinková and Prach, 2014). To accomplish this, drainage ditches must be blocked, either by human intervention or allowing natural succession, to raise the water table (Bragg, 2011). Following this, if restoration has been successful, recolonization of the area by peatland species will occur, eventually leading to C accumulation (Komulainen et al., 1998). Previous studies have demonstrated a return to a C sink by rewetted peatlands, although the length of time involved varies (Samaritani et al., 2011; Wilson et al., 2013; Knox et al., 2014;

Strack et al., 2014). Levels of CO2 uptake vary depending on how wet or dry the restored peatland is with wet restored sites having higher uptake rates due to less reduced peat decomposition. Wet restored peatland produces considerably higher amounts of CH4 compared to dry (sites retaining a deep water table) restored peatland (Strack et al., 2014). Vegetation responds slowly to changes in environmental conditions. While species die back occurs quickly, the establishment of species on a site may take a number of years (Couwenberg et al., 2011) which will influence the sink function of the peatland. In general, rewetted and natural bogs have low fluxes of N2O due to anoxic conditions (Byrne et al., 2004; Beetz et al.,

2013; Beyer and H. Höper, 2014). It is therefore hoped that by rewetting and restoring drained and damaged peatlands, the degraded peatland ecosytems will return to having a net cooling effect on the atmosphere.

Site topography may vary across a restoration site, influencing the water table rise (Ketcheson and Price). It is recommended that site managers level the site, reducing the variability of water table across the site (Strack et al., 2014). Previous land use will limit the ability of the site to act as a sink for C by the creation of microsites dryer than the surrounding areas. This recommendation is particularly prudent in forested peatland restoration as furrows and ridges, remaining from the initial ploughing of the site, vary greatly in wetness of peat, depth of peat and vegetation present.

Peatland restoration is a comparatively new area of research that has gained in popularity since the 1990s (Rochefort et al., 2003). Many projects across the globe have contributed to the present body of knowledge surrounding peatland restoration (Sliva & Pfadenhauer, 1999; Tuittila et al., 1999; Wilcox et al., 2006; Poschlod et al., 2007; Armstrong et al., 2009; Whinam et al., 2010; Mahmood & Strack 2011; Xiaohong et al., 2012). While the primary objective of peatland restoration is to repair the degraded ecosystem (Gorham and Rochefort, 2003), it may in fact not be possible to restore the peatland ecosystem to its original condition but restoration attempts to ensure a functioning C accumulating ecosystem is returned to the peatland (Vasander et al., 2003; Päivänen and Hänell, 2012). Restoration approaches differ between regions stemming from prior use, peat extraction methods and conventional thinking. In much of Europe there is the view that the long term aim of restoration on drained, degraded and damaged peatlands is the regeneration of vegetation of natural or undisturbed peatlands, dominated by Sphagnum mosses (Money and Wheeler, 1999) principally through rewetting. Peatland restoration in North America aims to restore peat producing vegetation communities of Sphagnum mosses, return the hydrologic cycle and productivity of peatlands (Rochefort, 2000). Following cessation of milling, turf cutting, peat extraction for horticulture or indeed forestry on a peatland, options for the further use of peatlands need to be considered. Many countries consider rewetting and restoration the best solution, thereby rejuvenating the peat accumulation process on the damaged peatland (Joosten, et al., 2012). Three routes exist for the restoration of peatland, (a) to allow the peatland to

37 recover spontaneously, (b) to direct the spontaneous recovery and succession towards a desired target and (c) to implement specific restoration measures (Konvalinková and Prach, 2014). Spontaneous recolonization of peatland by Sphagnum species may not be possible on most degraded sites where the acrotelm and therefore the seed source has been removed (Rochefort, 2000) and where restoration action is required. Trees grown after drainage are typically felled and removed. Blocking and damming drains will raise the water level within the first year; however the return of peatland species to the site may take a number of years to achieve (Päivänen and Hånell, 2012). Merely vacating the drained, degraded peatland without conducting any rewetting activities risks bog fires, further degradation, continued peat subsidence and erosion (Cris, et al., 2014). Conditions on rewetted sites vary in vegetation, hydrological functioning, past management, pH and length of drainage (Höper et al., 2008; Rydin and Jeglum 2013), all of which contributes to the success of returning to a C sink.

2.4.1 Rewetting and Restoration activities internationally

Internationally governments are realising the long term economic and environmental benefits of conserving healthy functioning peatlands as well as supporting and instigating efforts to restore what has been damaged (Cris et al., 2014). The continuing and well established threat of climate change remains a primary focus for governments worldwide and peatlands, with their carbon sink function, provide one option for reducing C emissions. Countries as far apart as China, Australia, Canada, Rwanda as well as the European countries of Poland, Belarus, United Kingdom, Sweden, Germany and Ireland have completed restoration projects adapted to their peatlands. Peatland restoration has been instigated for a number of reasons across the world (e.g. Tuittila et al., 2000a; Wild et al., 2001; Hope et al., 2005; Cris et al., 2013; Xiaohong et al., 2014). Since 1990, North American governments and the horticultural peatland industry have provided funding for restoration programmes (Rochefort, 2000; Cris, et al., 2014). Peatland restoration in North America aims to restore peat producing vegetation communities of Sphagnum mosses and return the hydrologic cycle and productivity of peatlands to their former condition (Rochefort, 2000). Canadian

Peatland restoration focuses on three activities- the reintroduction of live Sphagnum plants and diasporas in a density ratio of 1:10 (1m2 of Sphagnum diasporas over 10m2 of peatland), laying down a straw mulch cover over the Sphagnum layer and the blocking of the drainage system (Rochefort, 2000; Rochefort 2003; Waddington et al., 2003). Fertilizer may also be spread in the restoration area, encouraging the growth of vascular plants and Polytrichum moss; a nurse species of Sphagnum moss (Strack et al., 2014). Restoration conducted on the Bois-des-Bel peatland in Canada has taken a different approach to many restoration projects. Research there has included a paleoecological analysis; thereby tracking the vegetation history of the particular peatland. Evidence of black spruce as a native to the peatland has been found, therefore following restoration the growth of black spruce in conjunction with Sphagnum has been accepted as it once dominated the Bois-des-Bel landscape (Lavoie et al., 2001).

Much closer to home, the UK has demonstrated extensive peatland restoration on raised bogs in Lancashire, Cumbria and Northern Ireland in addition to blanket bog in areas such as the Flow Country in Scotland, Lake Vyrnwy and Migneint in Wales, the Peak District and Pennines (Bain et al., 2011). Specific afforested peatland restoration projects have occurred at Kielder Forest and Black Snib in northern England, followed by Langlands Moss (raised bog) in 1994 and Flow Country (blanket bog) since 1995, followed by several projects in Scotland, Wales and northern England (Bragg, 2011). However, most of these projects have had no formal monitoring and no research conducted on the success or failure of the work (Anderson, 2010). A policy and assessment procedure needs to be developed to assess the outcomes of restoration projects. Natural peatlands exhibit a unique environment and many British restoration projects have acknowledged the difficulty in achieving the exact former ecosystem of the peatland and instead have set different specific project aims (Anderson, 2001), such as maintaining and improving open mires, rather than regaining any specific vegetation community in the Border Mires Project (Burlton, 1996) and restarting peat formation, rather than re- establishing a particular vegetation type in the Stell Knowe restoration project (Stoneman, 1994; Longman, 1996). The Forestry Commission Wales have developed a field assessment tool to allow more accurate identification of forestry sites with the greatest restoration potential (Vanguelova et al., 2012). The use of this

39 tool requires access to detailed soil maps and visual assessment of peat cracking (Anderson et al., 2014) however it could potentially be adapted to suit other country’s needs.

Rewetting of peatland forestry was conducted as part of restoration projects like those in the Border Mires, the Caithness and Sutherland Peatlands or the Berwyn and South Clwyd Mountains SAC and the Migneint-Arenig-Dduallt SAC. In particular in the Border Mires tree removal, drain blocking by plywood, plastic sheeting, plastic piling, peat bunds and heather bales was demonstrated in conjunction with the installation of water level range gauges (WALRAGS) to monitor water table (European Commission, 2015). Experience gained by Anderson (2010) demonstrates that felling is essential for peatland restoration in Britain. Techniques which simply block drains have failed in Britain as the trees continue to grow in waterlogged conditions. Long term removal of regenerated trees has proven to be necessary until the water table becomes consistently high enough to suppress Sitka spruce regeneration. It remains unclear as to whether water table adjustment alone is enough for successful restoration; however the establishment of a more favourable peatland habitat is possible (Lunn and Burlton, 2010).

A review of best practise and techniques for peatland restoration, commissioned by ‘IUCN UK Peatland Programme‘s Commission of Inquiry on Peatlands’ has compiled common and best practise methods in the UK (Lunt et al., 2010). Government agencies in the UK have an active interest in biodiversity and peatland management; particularly peatland forestry, as evidenced by their plans and strategies’; i.e. Open Habitats Policy (Forestry Commission England, 2010), UK Forestry Standard (Forestry Commission, 2011), Biodiversity Strategy for England (DEFRA, 2011), Natural Environment White Paper (HM Government, 2011). The three categories of peatland restoration methods of water management, revegetation and vegetation management have been extensively tested in British restoration projects (Lunt et al., 2010). Guidance documents based on British restoration projects have been published from which those rewetting and restoring peatlands can seek guidance (Table 2.4).

Table 2.4 Peatland Restoration Guidance documents

Restoration Technique Guidance Document Ditch/ Grip blocking Holden (2009) A grip blocking overview. Report for the Environment Agency. Project 30254994. Armstrong et al (2009) Drain-blocking techniques on blanket peat: a framework for best practice. Journal of Environmental Management, 90, 3512-2519. Revegetation of Bare Anderson et al (2009) Moorland Restoration: potential and Peat progress. In Drivers of Environmental Change in the Uplands. (eds Bonn, A., Allott, T., Hubacek, K. and Stewart, J.), pp. 432-447. Gully Blocking Evans et al (2005) Understanding Gully Blocking in Deep Peat. Moors for the Future Report No. 4 Removal of scrub Brooks and Stoneman (1997) Conserving Bogs. The and forestry Management Handbook. Anderson (2010) Restoring afforested peat bogs: results of current research. Reintroduction of Carroll et al (2009) Moors for the Future Report No 16 Sphagnum Sphagnum in the Peak District Current Status and Potential for Restoration. Hinde et al (2011) Moors for the Future Report No 18 Sphagnum reintroduction project: A report on research into the re-introduction of Sphagnum mosses to degraded moorland Quinty and Rochefort (2003) Moorland Restoration Guide

2.4.2 Rewetting and Restoration activities in Ireland to date

In Ireland, restoration is used to encompass almost all rehabilitative and restorative work on peatlands. In Ireland to date, bog restoration works have been carried out by a number of private and state owned organisations; the Irish Peatland Conservation Council, National Parks and Wildlife Service (NPWS), Bord na Móna and Coillte. Coillte have managed three rewetting and restoration projects since 2002 funded by the EU LIFE Programme. A 5 year blanket bog restoration project was initiated in 2002. Restoration works were completed on 20 sites, totalling 1967 ha (Coillte, 2008). Sites were predominately located in western counties where the majority of Irish blanket peatland is located. Drawing on experience from England, Coillte adapted and trialled a number of restoration methods throughout the duration of this project.

(i) Fell to waste: On some sites; i.e., Garrane, Pollagoona, Carrick Barr, Eskeragh 1, Corravokeen and Cappaghoosh, trees were felled to waste using chainsaws. This has the effect of reducing the rate of recovery of ground vegetation as more of the area is covered by brash, thereby slowing restoration, furthermore, ground conditions are very difficult to traverse with the trees on the ground and drain blocks difficult to install. (ii) Fell and windrow: Following felling either by chainsaw or machine, trees were windrowed, i.e., piled in long lines using an excavator, thereby clearing large areas of the peatland surface allowing for greater regeneration of peatland species. In addition, further restoration work, drain blocking, dam establishment, water table monitoring, could be carried out with greater ease. (iii) Fell and chip: Originally proposed as a means of reducing the woody material remaining on the peat surface following felling, chipping was limited to Emlaghdauroe and Eskeragh 2. Vegetation regeneration has been successful where it was used but it was deemed too expensive, time consuming and limited by wet ground conditions to use on most sites, therefore it was not favoured over fell and windrow.

(iv) Installing dams: Coillte inserted plastic piling into drains as a means of installing dams on this project. In total over 63,300 dams were installed throughout the 5 year project.

Restoration of raised peatland has also occurred under LIFE funding on Coillte’s LIFE04 and LIFE09 projects, in association with the NPWS (http://www.npws.ie/researchprojects/wetlands/), utilizing the techniques agreed upon in LIFE02 (Coillte, 2005). Fresh organic material left on the peatland following felling has a number of potential impacts on CO2 emissions. As brash decays, CO2 is generated, contributing up to 40% of the CO2 losses from the peat (Mäkiranta et al., 2012). In addition, evidence for the ‘priming effect’ i.e. the increased decomposition of peat beneath the brash caused by the introduction of fresh organic material from the brash, has been demonstrated in previous studies (Fontaine et al., 2004; Fontaine et al., 2007). The introduction of fresh organic material provides microbes in the peat with a fresh energy source, allowing them to further decompose the peat (Mäkiranta et al., 2012). Brash material remaining on the site after clearfelling has been found to modify the soil nutrient content, particularly raising the potassium content (Palviainen et al., 2004; Hytönen and Moilanen, 2015). Increased nutrients will contribute to the priming effect, increasing microbial activity. Brash remaining on sites may also increase N2O fluxes (Mäkiranta et al., 2012). Given the potential effects for GHG emissions of brash remaining on the site, those rewetting forestry sites should endeavour to remove as much of the clear- felled material as possible prior to rewetting.

Initially Bord na Móna focused on conservation of undeveloped peatland. Prior to 1996, Bord na Móna had either transferred ownership or halted production plans on 5102 ha of peatland in efforts to conserve the integrity of the peatland (Bord na Móna, 1995). More recently, as production ceases on cutaways, focus has centred on rehabilitation and restoration of degraded peatland. Approximately 10,500 ha of peatland have been rehabilitated under Bord na Móna’s biodiversity projects (NPWS, 2013). Natural rehabilitation and succession has occurred in some areas while targeted drain blocking and dam building measures have ensured rewetting in other areas (NPWS, 2013). Some sites, due to deep peat extraction can no longer support peat forming Sphagnum species (Bord na Móna 2010a) and the return of their C sink function is uncertain (Wilson et al., 2007). Therefore, wetland areas and

43 natural woodland have been allowed to develop and succession continues (Renou- Wilson et al., 2010; Wilson et al., 2013). Bord na Móna have to date created in the region of 11,000 ha of wetlands on cutaway peatlands with up to 40,000 ha of cutaway becoming available for wetland creation in the coming years (Wilson et al., 2012). Rewetting coupled with restoration works have led to positive results (i.e. raised water table and re-establishment of vegetation) at Oweninnny Bogs, Co. Mayo, Abbeyleix (Killamuck) Bog and Ballycon Bog, Co Offaly (Bord na Móna 2010a; Irish Peat Society 2015). As production finishes on Bord na Móna cutaway, biodiversity projects will be established according to Bord na Mon’s biodiversity plans. Some of these plans will involve research trials such as the reed-bed establishment trial at Blackwater, West Offaly; Reed Canary Grass trial at Kilmacshane bog, East Galway; while others involve rehabilitation according to best practise (Bord na Móna 2010a). Bord na Móna cutaways have been utilised for renewable energy production with the establishment of windfarms on Mountlucas and Oweninny. Tourism and amenity are also recognised as having substantial potential as a land use for the Bord na Móna bogs in the future (Bord na Móna 2016).

The Irish Peatlands Conservation Council, based in Co Kildare, owns four peatlands nature reserves in counties Kildare, Meath and Waterford (Irish Peat Society 2015). Adopting the Canadian method of restoration as outlined in Quinty and Rochefort (2003), they have transferred Sphagnum moss from donor sites to bare cutover areas of their sites (O Corcora, 2015). This has been combined with drain blocking and tree removal (Irish Peat Conservation Council, 2014) in the hopes of restoring these peatlands to their natural function. So far the work appears to be a success. The water table has risen and remained steady on the site and three Sphagnum cutivation plots have been established on Girley Bog, Co. Meath (O Corcora, 2015).

Further options for Irish bog restoration should consider bog woodland creation or regeneration. Results from Germany indicate that undrained, natural bog forests act as long term modest carbon sinks (Hommeltenberg et al., 2014.)

2.4.3 Impact of rewetting and restoration on peatlands

Rewetting has been and remains the main method of restoration but it may in fact intensify water erosion on the peatland (Gorham and Rochefort, 2003). Agencies involved in restoration must evaluate the hydrology and peat structure of a particular site, drawing up management plans to counteract hydrological difficulties stemming from rewetting. The rate of decomposition and production on the damaged peatland prior to restoration will influence the pace of the return of the original carbon balance and ecological conditions (Waddington et al., 2003). Methane emissions resume following rewetting (Tuittila et al. 2000, Waddington and Day 2007, Couwenberg 2009, Wilson et al. 2009) although fluxes tend to differ from natural sites with nutrient rich sites emitting higher CH4 and nutrient poor sites lower CH4 than their undrained counterparts (Joosten, 2015). Peat oxidation rates are low after a site is rewetted due to the anoxic soil conditions and the majority of C sequestered by the peatland vegetation is stored in the peatland biomass pool (leaves, stems, roots). As time since rewetting increases, a decrease in the volume of CO2 sequestered annually has been reported (Yli-Petäys et al. 2007). Over time, the biomass pool of the peatland increases, approaching a steady state C sequestration point (Anderson et al. 2008). While a key aim of restoration or rewetting is to improve the C sink potential of peatlands, it is by no means a short term or predictable outcome. Greenhouse gas fluxes on rewetted bogs are primarily driven by soil temperature, water table and vegetation type, similar to their natural counterparts. Rewetting is often followed by a decrease in CO2 emissions and an increase in CH4 emissions while also cutting

N2O emissions (Tuittila et al., 2000b; Waddington and Price, 2000: Regina and

Myllys, 2009; Beyer and Höper, 2015). CO2 emissions fall as rising water table levels increases the depth of anoxic zone and greater volumes of vegetation on site are photosynthesising (Komulainen et al., 1999; Tuittila et al., 2000a; Dixon et al.,

2014). Where bare peat remains on rewetted sites, CO2 emissions will remain high (Dixon et al., 2014). The increased presence of vegetation like Eriophorum spp. acting as a conduct for gas transport and enhanced substrate availability contribute to heightened CH4 emissions (Waddington and Day, 2007). Areas surrounding blocked drains and ditches pose particular challenges as CH4 hotspots (Cooper et al., 2014).

However, in the long term, CH4 emissions are expected to fall and CO2 uptake will

45 increase as peat starts to accumulate (Augustin and Joosten, 2007). N2O emissions decrease following rewetting as plants out compete microbes for the available - nitrogen or nitrate (NO3 ) is fully reduced to nitrogen (N2) (Silvan et al. 2005, Glatzel et al. 2008, Roobroeck et al. 2010). Dissolved organic carbon concentration decreases following restoration returning to similar levels as those seen in natural peatlands (Glatzel et al., 2003). While rewetting reduces peat decomposition and the peatland may become a C sink, it may not in fact return to a condition where active peat formation is occurring. Active peat formation requires the re-colonisation of the site by peat forming species, such as Sphagnum which may not always be present. Several restoration projects have been the focus of post rewetting research particularly peatland harvesting sites (Marinier et al., 2004; Wilson et al., 2013; Beyer and Höper, 2015) and agricultural peatland (Tiemeyer et al., 2005; Kieckbusch and Schrautzer, 2007; Knox et al., 2014) while rewetted forestry remains understudied with few studies on any aspect of post peatland rewetting available (Komulainen et al., 1998; Komulainen et al., 1999; Haapalehto et al., 2011; Juottonen et al., 2012; Tarvainen et al., 2014) (Table 2.5). The results of studies following restoration have generally compared favourably with drained sites, proving their capability as sinks for carbon dioxide given the right conditions (Urbanová et al., 2012; Schrier-Uijl et al., 2014; Strack et al., 2014), however the strength of the sink function of restored sites varies with vegetation (Kivimäki et al., 2008).

-1 -1 Table 2.5 Carbon gas fluxes (CO2-C, CH4-C, g m yr ) from rewetted peat extraction, agricultural and forestry sites.

Land Use CO2-C CH4-C Years Authors g C m2 yr-1 g C m2 yr-1 since rewetting Peat 163–760 2.4–29.1 11 Wilson et al., Extraction (2007 & 2009)

111.1 2.3 10 Strack and Zuback, (2013) 93-304 ND between 10 and > 20 Wilson et al., 2015 ND 4.2 3 Waddington et al., (2007) −201.7 ± 126.8 – 16.2 ± 2.2 – >20 Beyer and Höper 29.7± 112.7 24.2 ± 5.0 (2015) -20 ± 5 ND 2 Waddington et al., (2010) Agriculture -397 39-53 02-May Knox et al., (2015) −446+±83 31.27±20.40 10 Hendricks et al., −311±58 32.27±21.08 10 (2007) −232±57 ND 10 Forestry ND 2.1 2 Komulainen et al., ND 4.6 2 (1998) 54 – 101 ND 2 Komulainen et al., 162 to 283 ND 2 (1999)

Positive flux values indicate either carbon emissions to the atmosphere (CO2 and CH4). Negative values indicate carbon removal from the atmosphere by the peatland. Where sources used a different convention it has been changed. ND denotes not determined.

Many restoration attempts have been too short to conclude whether the original ecosystem will be restored (Gorham and Rochefort, 2003). However it has been estimated that a significant coverage (20- 60%) of bog vegetation can be re- established in a peatland 3-5 years post restoration, a constant high water table maintained after ten years and a peat forming, fully functioning ecosystem established after 30 years (Rochefort, 2003) although the last step may take longer. Studies on rewetted peatlands have demonstrated little consistency in their GHG dynamics and have proven to be unpredictable and site specific. A rapid return of the peatland C sink function has been reported in some studies (Wilson et al., 2011;

Urbanová et al., 2013); while in others high CO2 emissions have persisted years after rewetting (Wilson et al., 2007b; Samartitani et al., 2011). The GHG dynamics are affected by a number of factors; time since rewetting, pH, nutrient status, peat type, vegetation present, hydrology which implies a tailor made approach is required for each rewetting project (Wilson et al., 2016a).

Candidate peatlands for restoration projects often exist with no exact histories of the individual ecosystem properties; therefore the restoration goal becomes more general. The level of disturbance on many candidate sites renders it impossible to determine the original state or particulars of the ecosystem, unless paleoecological studies are conducted (Gorham and Rochefort, 2003). Without monitoring data prior to restoration, it is difficult to evaluate the success of the work (Lunt et al., 2010). Restoration of afforested peatland is an area of research lacking in a database of peer reviewed literature. Lack of literature in the area leads to key questions for which answers are not available related to drain blocking, tree felling, using identical restoration techniques on both raised and blanket peatlands (Bragg, 2011). Restoration treatment results in England suggest that felling to waste rather than whole tree removal or conventional harvesting methods maintains a higher water table on the peatland, even during prolonged summer dry spells (Anderson, 2010). In order to reduce evapotranspiration on the peatland, a number of measures during restoration projects are used; clearfelling, complete removal of the trees and brash and planting a wind shelter on marginal areas (Schumann and Joosten, 2008). Evidence from British projects confirms that trees must be felled to ensure rewetting and restoration (Anderson 2010), therefore given the similarities between the Irish and British climate and forestry, it can be assumed that the most successful restoration projects will always include tree felling. Research from a laboratory in

Wales has suggested that the drain blocking method influences the level of CH4 emissions from rewetted peatland (Green et al., 2014), although field experiments would be needed to quantify the effects of methods used in Ireland.

While data on varying aspects of the relationship between CO2 and CH4 emissions, biotic and abiotic factors and management exist for rewetted peat extraction sites both in Ireland and further afield (Cobbaert et al., 2004; Wilson et al., 2009; Kollmann and Rasmussen 2012; Urbanová et al., 2013), data gaps for agricultural restoration and rewetted forest remain. Research on agricultural restoration has occurred in the USA (Herbst et al., 2013; Knox et al., 2014) and in small pockets across Europe (Bakker and Olff 1995; Hendriks et al., 2007; Niedermeier and Robinson 2007; Graves and Morris 2013). Published research results on greenhouse gas balances in rewetted forestry sites are greatly lacking with very few studies available (Komulainen et al., 1999; Koskinen et al., 2016). Results booklets, reports

48 on best practise and reasons for such restoration are available (Anderson 2001; Coillte 2008a; Coillte 2008b; Anderson 2010; Klemedtsson and Rova, 2015). Increased attention from the scientific community to rewetted forestry is required to fill the information gap, provide emission factors to policy makers and allow land owners to make decisions based on proven research.

2.4.4 Methods of determining success or failure of rewetting and restoration

The definitive test of success in a restoration project is whether the restoration goals and objectives have been achieved (Wheeler and Shaw, 1995). Setting appropriate goals therefore becomes a vital element of the restoration process (Kentula 2000; Hobbs and Harris, 2001). A return of the site to its original or ‘natural’ state may be unrealistic as climate and environmental change in addition to anthropogenic influence may make it impossible to re-establish this state. Therefore specific site conditions or ecosystem developmental stages as restoration goals may be unattainable. An alternative to this is to establish peatland services or characteristics as goals; i.e. structure or function of the peatland (Bonnett et al., 2009). Specific factors which can be assessed are hydrology, vegetation, birds, invertebrates and growth of peat (Wheeler and Shaw, 1995). Objectives of restoration projects need to be clearly established prior to any onsite monitoring whether it be pre or post restoration works (Bonnett et al., 2009). It has been argued, in cutover sites in particular, that successful restoration is only achieved once sufficient fresh organic matter has accumulated to maintain the mean water table of a drought year above the underlying cutover peat surface (Luchesse et al., 2010). A detailed set of monitoring guidelines providing information on peat bulk density, humification, greenhouse gas fluxes, water chemistry vegetation, birds, microorganisms and more has been compiled by Bonnet et al., (2011), providing a comprehensive guide to restoration practitioners.

Ideally monitoring on the proposed restoration site should occur before and after restoration works to provide a baseline from which to track changes or failing that, comparisons with a carefully assessed or control site (Clewell et al., 2005; Bonnett et al., 2009). Monitoring is vital in restoration projects as all projects should be viewed

49 as an opportunity to gather information on techniques, provide data on the conditions prior to restoration works thereby allowing present and future site managers to track changes and progress throughout the process (Wheeler and Shaw, 1995). The prediction of future restoration outcomes using early monitoring can allow evaluation on the necessity of additional work to correct undesired successional outcomes (González et al., 2014). Long term monitoring of restoration sites may be required to track changes which may take decades to occur (Bonnett et al., 2009). Indicators are often used to evaluate the progress of a restoration site. Using indicators could introduce bias into the appraisal of the restoration success as they assess the condition and progression of an ecosystem based on basic estimators, i.e. the occurrence of a specific species (Dale and Beyeler, 2001). González et al., (2013) found that while using plant species identified as significant indicators of successful restoration, differences in frequency and cover of these species varied little between different categories of restoration outcomes, creating difficulty in predicting restoration certainty. While monitoring plant species in a restoration project is an essential activity, it alone is not a sufficient predictor of long term success (Herrick et al., 2006). Vegetation and other abiotic factors, i.e. hydrology, nutrient cycling or carbon exchange, should be monitored to assess restoration success (Zerbe et al., 2013). Monitoring techniques used post restoration works, must be related to the project’s goals and objectives; avoid the use of indicators in assessing specific results, i.e. a reduction in greenhouse gas emissions (Bonnett et al., 2009).

2.5 Impacts of drainage and land use change and rewetting on peat properties

2.5.1 Bulk Density

Bulk density indicates the level of soil compaction depending on the anthropogenic impact which is vital in understanding nutrient budgets (Hossain et al., 2015). Natural peatlands display dry bulk densities ranging from 0.066- 0.112 g cm−3 (Byrne et al., 2004). Bulk density of peat increases following drainage, as the physical structure of the peat becomes more compact through continued

50 decomposition (Minkkinen and Laine, 1998; Päivänen and Hånell, 2012) (Table 2.6). Peat bulk density may be used as an indirect indicator of the degree of decomposition (Boelter, 1969). Subsidence can be quickly recognised as an increase in bulk density in the peat profile. Subsidence of the peat above the water table can be calculated by contrasting the bulk densities of the peat above and below the water table (Schothorst, 1997). Bulk density values for peatlands under agriculture are higher (Okruszko and Ilnicki, 2003), with values up to 0.95 g cm-3 reported (Leifeld et al., 2011b). Forest drainage on peat also has a significant impact on bulk density, increasing it from 0.082 g cm-3 to 0.133 g cm-3 in Finland (Minkkinen and Laine, 1998). Some studies have found that bulk density of peat increases with depth (Howard et al., 1994; Milne and Brown, 1997; Cruickshank et al., 1998), however, more recent studies have revealed that in some cases bulk density may show no change (Tomlinson and Davidson, 2000; Lewis et al., 2011) or even decrease with depth (Weiss et al., 2002). Raised bogs have been found to have significantly higher bulk densities than blanket bogs (Wellock et al., 2011). Sites following rewetting maintain their high bulk densities (Wilson et al., 2008; Zak et al., 2010), although in some instances there is a slight reduction compared to drained sites (Tuittila et al., 2000a).

Table 2.6 Published bulk density values from natural peatland and drained peatland forestry

Land use Bulk density (g cm-3) Reference Natural ~0.05-0.11 Minkkinen and Laine (1998) ~0.05-0.13, 0.03-0.08, 0.05-0.1 Anderson (2002) 0.01-0.1 Leifeld et al., (2011a) Drained forested ~0.10-0.16 Minkkinen and Laine (1998) 0.4-0.14 Leifeld et al., (2011a) 0.089-0.127 Wüst-Galley et al., (2016)

2.5.2 Ash content

Peat is defined as accumulated matter consisting of no less than 30% dead organic material (Renou-Wilson et al., 2011). The material which remains following ignition of a peat sample is known as the soils ash content. Ash content is positively linked to bulk density (Leifeld et al., 2011a). Organic soils naturally have low ash content but

51 a rise in atmospheric deposition and drainage encouraging organic matter decomposition will increase ash content (Leifeld et al., 2011a). Therefore drained peatlands have greater ash content than natural sites. Carbon losses since the onset of drainage can be calculated using the ash content changes with depth in a peat profile under a number of assumptions, (a) ash content was the same in all depths prior to drainage, (b) the upper layers started to oxidise following drainage, (c) oxidised C is released as CO2 from the peat surface and (d) ash content remains the same as before drainage in the lower peat layers (Rogiers et al., 2008). Although Leifeld et al., (2011 a) determined that an increase in ash content throughout the profile is not uniquely caused by drainage, it was concluded that using the ash residue method is an appropriate method for calculating carbon losses from bogs. The effect of ash content on peatland respiration is as yet unclear. Studies carried out by Hogg et al., (1992) found variable results with increased ash content promoting respiration in many of the test samples but not all.

2.6.1 International policy on peatlands

Many international conventions, agencies and agreements influence peatland conservation and management either directly or indirectly such as, the convention on biological diversity, united nations framework convention on climate change, food and agriculture organization of the United Nations, intergovernmental panel on climate change and the Ramsar convention (International Peat Society 2014). The role of wetlands in our environment, as important ecosystems and habitat, has been recognised by policy makers around the globe for many years with the Ramsar convention existing since 1971 (Matthews et al., 1993). The European Union maintains a union wide network of protected sites, which includes specific peatland sites as special areas of conservation (Joint Nature Conservation Committee, 2010). Peatland restoration is eligible for funding from the EU Life programme and the common agricultural policy through the financial support of agri-environmental schemes allows for sustainable management of peatland while sustaining livelihoods (Cris et al., 2011). In recent years there has been an increasing focus on the role rewetted or restored peatlands have to play in carbon emission or sequestration and carbon accounting.

The 2006 IPCC Guidelines for National Greenhouse Gas Inventories (Eggleston et al., 2006) do not include guidance on rewetted peatlands highlighting both the lack of research conducted and the level of priority given to such sites at that time. However, the 2013 Supplement to the 2006 IPCC Guidelines for National Greenhouse Gas Inventories: Wetlands (IPCC 2014b), dedicates a chapter to rewetted organic soils demonstrating the shift in a few short years. Tier 1 emission −1 −1 factors (EFs) for CO2 in temperate rewetted soils are -0.23 ±0.41 t CO2-C ha yr −1 −1 (rich) and + 0.50 ± 1.21 t CO2-C ha yr (poor) where negative values indicate removal of CO2 from the atmosphere. Despite peatlands sharing similar characteristics and undergoing similar drainage and management treatment in many cases, emission levels vary considerably across studies (Turetsky et al., 2014), even within the same geographical zone. The IPCC wetlands supplement (IPCC 2004b) attempts to encompass the whole temperate zone into one EF for each individual land use, thereby simplifying the reporting process, in efforts to encourage countries to report emissions (IPCC 2004b). The lack of data for individual peatlands is also responsible for this method. Country specific as well as management specific factors need to be developed to improve accuracy and reliably in the system. Wilson et al., (2015) argues that Irish emissions are most likely 40% lower than those calculated using the prescribed IPCC Wetlands supplement (IPCC 2004b) values, as demonstrated by their focused study on Irish and Scottish peat extraction sites. Ireland requires an increased research presence on peatlands in order to develop county specific emission factors. Research on peatlands has to date focused on natural (Laine et al., 2007a; Laine et al., 2007b; Koehler et al., 2009; Sottocornola and Kiely 2010) and rewetted industrial peatlands (Farrell and Doyle 2003; Wilson et al., 2013). Results from drained peat sites originate in other European countries or further afield (Glatzel et al., 2003; Hyvönen et al., 2009; Salm et al., 2012). Information on Irish agricultural drained peatlands and rewetted forestry remains absent. Following restoration work by Coillte, it is important that emissions are quantified for rewetted peatland forests. This subject is the focus of this project which aims to provide emission factors for that land use. This will increase the scientific community’s knowledge of our national emissions and provide greater accuracy in Irish carbon accounting.

Chapter Three

Carbon dioxide flux dynamics in a blanket peatland forest eight years after rewetting

3.1 Abstract Forest drained peatlands in Ireland are currently undergoing rewetting in order to return these ecosystems closer to their natural condition. Drained peatland soils are significant sources of carbon dioxide (CO2). Rewetting is considered to be an important climate change mitigation tool and a means of creating conditions for carbon (C) sequestration. The aim of this study was to investigate the controls on gross photosynthesis (PG), ecosystem respiration (RTOT) and net ecosystem CO2 exchange (NEE) and assess whether the C sink function of the peatland had been restored 8 years following rewetting. Since rewetting, water table levels have risen on the study site and more typical peatland vegetation species such as Molinia caruela, Sphagnum spp, Calluna vulgaris, Potentilla erectus, Eriophorum angustifolium, Campylopus spp, Polytrichum commune and Hypnum jutlandicum have colonised the site. In this study, CO2 fluxes were measured on a rewetted blanket peatland using the chamber method. Microsites comprising the dominant vegetation on the study site were established and measured for one year. Both the -2 -1 Eriophrium vaginatum (184.8 ± 82.8 g CO2-C m yr ) and Molinia plots (168.4 ± -2 -1 229.5 g CO2-C m yr ) were CO2-C sources during the study year, however the Calluna vulgaris- Cladonia portentosa plot acted as a modest C sink (-142.84 g -2 -1 CO2-C m yr ). Although water table was not used in all models, annual losses generally followed changes in water table. The greatest C losses were from the driest microsites. Temporal variation was most obvious in gross photosynthesis (PG) which decreased substantially in the winter months as leaf area and photosynthetic photon flux density (PPFD) are at their lowest. The annual CO2–C balance was 131.6 ± 298.3 g m-2 year-1. Our results show that C losses on this site are much lower than from drained peatland sites and CO2 fluxes are largely determined by soil temperature, water table and vegetation.

3.2 Introduction

Peatlands in Europe are estimated to cover 329,000 km2 (Montanarella et al., 2006), storing approximately 42 Gt of C (Joosten and Clarke, 2002). In Ireland, blanket peatland accounts for 12 % of the land area and stores an estimated 0.57 Gg of C or 28 % of the country’s soil C stock (Tomlinson, 2005). Blanket peatland is considered to be an important but scarce habitat (Conaghan, 2000). It differs from other peatland types in its location, morphology, depth, dominant vegetation and rainfall required for formation (Foss and O’ Connell, 1996). The C sink function is driven by plants primary productivity remaining greater than the impeded rate of decomposition in the acidic, water logged, and cold peatlands (Renou-Wilson et al., 2011; Päivänen and Hånell, 2012). Peatlands are highly susceptible to climatic and hydrological changes which can switch their role from sink to source of C to the atmosphere (Strack, 2008a). Peatlands occupy an important position in mitigating climate change due both to their current C stock and future sink capacity (Yu et al., 2011). Peatlands are primarily affected by anthropogenic actions, particularly drainage for forestry, agriculture or peat extraction (Waddington and Price 2000; Minkkinen et al., 2007;

Salm et al., 2012), turning them into long term sources of CO2 (Silvola et al., 1996; Ojanen et al., 2010). Drainage for forestry influences fundamental ecosystem features such as microbial activity, aeration of the surface peat and nutrient conditions (Laine et al., 2006), increases soil bulk density and causes peat subsidence (Minkkinen and Laine et al., 1998) and triggers the replacement of peatland vegetation by woodland species (Korpela 1999; Laine and Vanha- Majamaa 1992). In Ireland, the area of peatland which is forested is estimated at approximately 322,000 ha (NFI 2013), much of it on blanket peatland (Coillte 2003). Some drained peatland sites have proved unsuitable for forestry and rewetting has been utilized as an alternative land use on these priority habitats for conservation. Rewetting is seen as a viable means of restoring the carbon sink function of peatlands (Höper et al., 2008). Rewetting is defined here as ‘the deliberate action of raising the water table on drained soils to re-establish water saturated conditions’ (IPCC, 2014b). Through drain blocking to raise the water table (Bragg, 2011) followed by a host of possible management measures devised for each individual peatland site (Quinty and Rochefort, 2003; Wheeler, 1995), rewetting aims to limit

58 further peat degradation (Schuman and Joosten, 2008) and re-establish a peatland ecosystem similar to the one which was degraded (Konvalinková and Prach, 2014). Previous studies on rewetted peatlands have demonstrated little consistency in their GHG dynamics and have proven to be unpredictable and site specific. A rapid return of the peatland C sink function has been reported in some studies (Wilson et al.,

2011; Urbanová et al., 2013); while in others high CO2 emissions have persisted years after rewetting (Wilson et al., 2007b; Samartitani et al., 2011). The GHG dynamics are affected by a number of factors; time since rewetting, pH, nutrient status, peat type, vegetation present, hydrology which implies a tailor made approach is required for each rewetting project (Wilson et al., 2016a), such as through drain blocking method selection or deciding whether to reintroduce Sphagnum spp. The initial conditions at rewetting also vary in terms of vegetation, hydrological functioning, past management, pH and length of drainage (Höper et al., 2008; Rydin and Jeglum 2013), all of which impacts on their trajectory in returning to a C sink. Even many years post rewetting, water table dynamics may not behave in the same way as natural blanket peatland (Holden et al., 2011), posing challenges for peatland restoration. The objectives of this study are as follows: (a) identify the controlling variables of

CO2 fluxes on rewetted peatland forest sites; (b) to estimate the annual CO2 balances for a range of microsites at a rewetted blanket peatland forest and (c) to compare our findings with other rewetted peatland sites and suggested management options for increasing the CO2 sink potential.

3.3 Materials and Methods

3.3.1 Site Description

Pollagoona is a lowland blanket bog (Conaghan, 2000), located in north County Clare (53°00' N, 8°32' W). It is located in the Slieve Aughty Mountains at an altitude of 156m above sea level, in a shallow saddle on gently sloping terrain (Fig 3.1). The site lies about 20 kilometres southeast of the town of Gort, Co. Galway and 11 kilometres southwest of Woodford, Co Galway. The climate is temperate and humid with cool summers and mild winters. The 30 year mean annual air temperature was 10°C (Met Éireann 1981-2010, Athenry Meteorological Station

(http://www.met.ie/climate/monthly-data.asp?Num=1875) and 30 year mean annual rainfall was 1435 mm (Met Éireann, 1984- 2015, average at Gort rainfall station). As Gort is closest to Pollagoona, rainfall data recorded there was used for annual means (https://drive.google.com/file/d/0B_bgNurSrUQdUHJ5a3hOQ3JUR2M/view?usp=sharing).

Figure 3.1 Location of the Pollagoona bog site

Figure 3.2 Site map of Pollagoona rewetted area (outlined in red) containing both forested and undrained peatland.

The site contains both natural bog and previously forested peatland. The restoration site totalled 60.6 ha of which 40.3 ha was forested; predominately with Sitka Spruce (Picea sitchensis [Bong] Carr) in the late 1980s. The main restoration activities which took place were manual felling to waste of conifers, windrowing in places and the blocking of man-made drains. Conifers were felled in 2006 (unpublished Coillte data) with drain blocking occurring soon after. Ridges and shallow furrows remain on the rewetted peatland. The vegetation present on Pollagoona is typical of lowland blanket bogs as outlined by Conaghan (2000), where species such as Molinia caerulea dominate, mosses (Polytrichum commune and Hypnum cupressiforme) are the next dominant species and Calluna vulgaris, Potentilla erectus and Eriophorum angustifolium occur in small quantities (See Fig 3.3). Eriophrium vaginatum was also present, although in less abundance than on Scohaboy, a raised bog site. Moss cover varies across the site with water table depth. Plots with a higher water table tended to contain more Sphagnum spp. such as Sphagnum capillifolium, Sphagnum papillosum, as well as non-Sphagnum spp., e.g. Campylopus species, Polytrichum commune and Hypnum jutlandicum than plots where the water table remained below 20 cm for much of the growing season.

3.3.2 Study Site

The site was established in January 2014, two months prior to commencing gas measurements. Eight steel collars (60 cm × 60 cm × 30 cm) were inserted into the peat in a line, running from East to West. This design was used to capture the observed hydrologic variation present on the site (observed by water table sampling and vegetation present) and vegetation variation between the furrows and ridges and the brash lanes. Plots were labelled P1- P8, P1 located at the wettest point on the line as water table depth decreased along the plots (Fig 3.3). P4 and P8 were situated in what remained of a brash lane. In summer, vegetation in 6 of the plots (P2, P3, P4, P6, P7 and P8) was dominated by Molinia caerulea and so were designated as Molinia microsites. Mosses (Polytricium commune, Sphagnum papillosum, Sphagnum capillifolium, Hypnum jutlandicum, Rhytidiadelphus spp., Racomitrium lanuginosum and Campylopus spp.) were also found in sample plots all year round. Given that P1 displayed different vegetation to other sample plots (Eriophrium vaginatum, Polytricium, Sphagnum papillosum, Eriophorum angustifolium and lower volumes of Molinia caerulea), it was designated as the ‘Eriophrium- Sphagnum’ microsite and P5, containing Calluna vulgaris, Cladonia portentosa, Sphagnum capillifolium, Eriophorum vaginatum and Potentilla erectus in addition to Molina caerulea, was deemed the Cladonia- Calluna microsite.

Table 3.1 Vegetation species recorded in the study plots. Species are listed in descending order of dominance.

Eriophorum-Sphagnum Cladonia- Mosses (P5) Molinia (P1) (P2, P3, P4, P6, P7, P8) Eriophorum vaginatum Molinia caerulea Molinia caerulea Polytrichum commune Sphagnum capillifolium Polytrichum commune Sphagnum papillosum Calluna vulgaris Hypnum jutlandicum Molinia caerulea Cladonia portentosa Sphagnum capillifolium Eriophorum angustifolium Eriophorum vaginatum Rhytidiadelphus spp. Potentilla erectus Racomitrium lanuginosum Sphagnum papillosum Campylopus spp. Juncus effusus

Figure 3.3 Pollagoona microsites in May 2014, prior to summertime Molinia growth. Top from left to right, P1 to P4, bottom from left to right, P5 to P8.

The collars were topped with a 3cm deep channel that was filled with water to provide an airtight seal during measurements. Boardwalks were constructed beside each collar to minimise disturbance and damage to peat surface and plant cover during measurements.

3.3.3 CO2 flux measurements

Light and dark chambers were used to measure instantaneous net ecosystem exchange (NEE). Net ecosystem exchange was measured using a static polycarbonate chamber (60 cm × 60 cm × 30 cm) fitted with a battery operated fan, whose function was to mix the air within the chamber head-space. The chamber was connected to a cooling system which maintained the internal temperature to within 1ºC of the ambient temperature. Carbon dioxide fluxes were measured using a portable infrared gas analyser (EGM- 4, PP Systems, UK). In order to measure instantaneous NEE, CO2 concentration initially was measured at 15 s intervals over a period of 60 s to 180 s under ambient illumination. The chamber was vented for a short time following the initial measurement of CO2 at each collar. To better establish a relationship between photosynthetic photon light density (PPFD) and photosynthesis, artificial shades were used to obtain fluxes under a greater range of PPFD. Following light measurements, the chamber was then replaced on the collar and covered with an opaque material and CO2 fluxes measured to determine total ecosystem (heterotrophic and autotrophic) respiration (RTOT) (method as per Wilson et al 2007, Tuittila et al., 1999 and Tuittila et al., 2004). Carbon dioxide flux rates were calculated from the linear change in gas concentration as a function of time. The ecological sign convention, in which fluxes from the atmosphere to the biosphere are negative, was used. Gross photosynthesis

(PG) was estimated as the sum of flux rate values measured in light (NEE) and dark

(RTOT). Carbon flux measurements were conducted on a fortnightly basis except for a one month period in December 2014-January 2015, from August 2014 to September 2015. Measurements were made between 8.00 am and 5.00 p.m. in a random pattern. Two measurement rounds were conducted on most sampling days.

PPFD sensor

Raising Collar

Cooling system

Thermometer

EGM- 4

Figure 3.4 CO2 flux chamber and raising collar, equipped with a cooling system and fan. An infrared gas analyser (EGM- 4) measures CO2 concentration. PPFD and air temperature are measured inside the chamber.

3.3.4 Environmental Variables

In order to correlate measured fluxes of gases to environmental conditions, air temperatures inside the chamber and in the peat at 5 cm and 10 cm (T5, T10) depths were measured simultaneously at sampling time. Soil temperatures were measured using a temperature probe (Jenway 220). Air temperature was measured using a Ted Pella Inc., 28163 Traceable Total-Range Thermometer. The level of the water table was measured using PVC pipes with a series of small holes pierced at regular intervals along the side. These pipes were positioned beside the collars prior to the start of the study, leaving 10cm of piping visible above the ground. Water table depth was measured using a Pocket Dipmeter Kill Mini 10 m in length supplied by HYDROKIT Dorset. The water table data was interpolated between measurements in order to describe water table patterns across the seasons. A meteorological station

(Watchdog 1400 Micro station) was installed on each site measuring PAR and soil temperature at 5 cm and 10 cm at ten minute intervals throughout the duration of the study. Measurements of all environmental variables began in March 2014.

3.3.5 Green Area Index

Green area index (GAI), which represents the green plant material, was used in gas flux modelling to relate gas fluxes to the seasonal changes in plant growth and senescence. The GAI of each sample plot was calculated based on Wilson, et al., 2007. Each sample plot was placed in vegetation communities’ representative of those present across the restored peatland. Within each site, five sub sample plots (8cm × 8cm) were established, one plot in each corner and one in the middle of the collar. Therefore the vegetation range in the microsite was captured. At fortnightly intervals, the total number of leaves and stems of individual plant species within the subsample plots were counted and used to estimate the average number of leaves and stems of each species per m2. In order to avoid disturbance to the plants inside the sample plots, 3 individual plants of each species outside the sample plots, were tagged and their leaves measured. These tagged plants were similar in age and stature to those inside the sample plots. The length and width of each leaf was measured and used to estimate leaf area. The area of the leaves was estimated using species-specific formulae, based on the geometric shape of the leaf (e.g. ellipse, circle, rectangle). These measurements were made on the same day as leaves were counted. If present, the green surface area of the plant stem was calculated using the formula for surface area of a cylinder; Green stem area= (2*π*r)*h (Eq. 3.1) where r is the radius and h is the plant stem height. A percentage vegetation cover was also conducted at the same time on all subplots. The GAI of plants with irregular growth patterns or irregular leaf edges was estimated from this percentage cover.

The average stem area (SA) and leaf area (LA) were calculated to give a representative leaf and stem size for each species on each sample day throughout the study period. The green area of each species, i (m2 m-2) was calculated using the following formula;

GAi = (LA * Ln) + (SA * Sn) (Eq. 3.2)

2 where LA is the average leaf area of the three tagged plants (m ), Ln is the average 2 2 leaf area per m , SA is the average stem area of the tagged plants (m ) and Sn is the average number of stems per m2.

3.3.6 Modelling the annual development of leaf area index

In order to describe the seasonal development of vegetation in the sample plots, plot specific model curves were applied. For all plots, the seasonal development of LAI was unimodal and normally distributed (Gaussian);

푗푢푙푖푎푛−푥 (0.5( 푚푎푥)2) Daily GAI: y+ GAImax e 푏 , (Eq. 3.3)

Where Julian is Julian day, GAImax denotes the maximal GAI of the plot during the season, xmax is the Julian day when maximum GAI occurs and y denotes the base of the curve b determines the curves shape.

3.3.8 Data Analysis

Similarly to Beyer and Höper, (2014), other influencing factors like water table or vegetation were only considered on a long term (i.e. fortnightly or monthly), not on a short term (i.e. daily), disregarding that these factors may also change in the course of the day. Therefore they were only recorded on sampling days. Missing water table data between sample days was interpolated for use in regression modelling.

3.3.9 Modelling of CO2 flux

Each of the CO2 components, RTOT and PG were modelled separately as they have different relationships with the environmental variables measured and in addition, we wanted to quantify the components individually. Gross photosynthesis was calculated from the addition of NEE and RTOT. Measured RTOT and calculated PG were then used in model construction. Gross photosynthesis is principally controlled by irradiation (PPFD), explained by a Michaelis- Menten hyperbolic function

(Stryer, 1988). Non- linear regression has been widely used in modelling PG, owing to the use of the hyperbolic relationship (Alm et al., 2007; Tuittila et al., 2004;

Wilson et al., 2007; Urbanová et al., 2012). PG usually has distinctive responses to the remaining measured variables: GAI- linear, temperature- exponential and WT-

Gaussian (Tuittila et al., 2004; Laine et al., 2009). RTOT is the sum of autotrophic plant respiration and heterotrophic soil respiration; the two processes were measured together in this study. RTOT is primarily linked to temperature, however WT and leaf area also influence the rate of fluxes. An exponential function is often applied to describe the relationship between temperature and RTOT (Lloyd and Taylor, 1994). Soil respiration has a saturating relationship with WT; respiration increases as WT falls until a particular depth at which point the surface becomes too dry and respiration is reduced (Tuittila et al., 2004). Plant respiration is linked to the metabolic activity of the plants and has a comparable response as WT. In this study,

GAI was not used in RTOT modelling as no relationship was determined. Nonlinear regression models were initially constructed separately for each individual collar and in instances where vegetation characteristics and controlling variables were similar, plots were grouped into microsites based on vegetation characteristics The models were then parameterized for each vegetation community as each differed in response to environmental variables. SPSS 22 for Windows statistical package (SPSS, Inc.) was used in flux modelling.Soil temperature at 5 cm, GAI and PPFD were used as explanatory variables for the PG models. Green area index was used in the CO2 flux models to explain the changes in CO2 flux which were related to the seasonal pattern of plant growth. Green area index was added to the model as either a linear term (Eq. 3.5) or an exponential rise to maximum (Eq.3.6). Temperature was added to the model as a linear (Eq. 3.4 and 3.5) or Gaussian (Eq. 3.6) term. Models were adapted from Wilson et al. (2015), (Eq. 3.4 and 3.5) and Laine et al. (2009a), (Eq. 3.6). Only variables which increased the explanatory power of the model were accepted. Water table did not improve the models and so was excluded. Model acceptance was based on the following criteria; (a) statistically significant model parameters (p<0.005), (b) highest possible coefficient determination and (c) lowest possible parameter error. The PG models used were as follows;

푃푃퐹퐷 PG= Pmax ( ) ∗ 푇5푐푚 (3.4) 푃푃퐹퐷+푘푃푃퐹퐷

푃푃퐹퐷 PG= Pmax ( ) ∗ 푇5푐푚 ∗ 퐺퐴퐼 (3.5) 푃푃퐹퐷+푘푃푃퐹퐷

2 푃푃퐹퐷 푇5푘푒푙 −푇표푝푡 PG= Pmax ( ) ∗ 퐸푋푃[−0.5( ) ] ∗ [1 − 퐸푋푃(−퐸 ∗ 퐺퐴퐼)] (3.6) 푃푃퐹퐷+푘푃푃퐹퐷 푇푡표푙

In equations 1-3 PG is gross photosynthesis, Pmax is maximum photosynthesis, PPFD is photosynthetic photon flux density, kPPFD is a fitted model parameter and denotes the value at which PG reaches half its maximum, GAI is green area index, T5cm is soil temperature at 5cm (°C), T5kel is soil temperature at 5cm (K) and Topt is the temperature optimum and Ttol is the temperature tolerance for photosynthesis and E is a fitted model parameter. The PG model parameters for each microsite are given in Table 3.2.

2 Table 3.2 Estimated parameter values and goodness of fit (r ) for PG. Number of the equation used is given for each microsite. Plots are labelled P1- P8.

Sample Plot PG

2 Pmax kPPFD Topt Ttol E R Eq. Eriophorum- Sphagnum P1 247.92 1026.72 0.785 3.4

Calluna- Cladonia P5 133.72 388.99 0.839 3.5

Molinia caerulea 3513.37 422.04 290.34 3.95 2.00 0.678 3.6 P2,3,4,6,7,8

To model soil respiration, RTOT, was linked to soil temperature at 5 cm (Kelvin) using the exponential relationship described by Lloyd and Taylor (1994). Neither water table nor GAI improved the models for the Eriophorum- Sphagnum or Calluna- Cladonia plots and so were excluded. Total respiration in these plots may be dominated by hetertrohic respiration as GAI did not influence the model. The Molinia vegetation communities displayed a quadratic relationship with WT. Sample plots were grouped and parametrised based on vegetation due to each vegetation community behaving distinctly from the other.

1 1 RTOT= [퐴 ∗ 퐸푋푃[B ( − )] (Eq. 3.7) 푇푟푒푓−푇0 푇− 푇0

1 1 2 RTOT= [퐴 ∗ 퐸푋푃[B ( − )] + (푊푇 ∗ 퐶) (Eq. 3.8) 푇푟푒푓−푇0 푇− 푇0 where Tref was set at 283.15K and T0 was set at 227.13K. The temperature minimum at which respiration reaches zero is denoted by T0, A, B and C are fitted model parameters. Air temperature (T) is given in degrees Kelvin. The RTOT model parameters for each microsite are given in Table 3.3.

2 Table 3.3 Estimated parameter values and goodness of fit (r ) for RTOT. Number of the equation used is given for each microsite. Plots are labelled P1- P8.

Sample Plot RTOT A B C R2 Eq. Eriophorum- Sphagnum P1 344.83 441.22 0.628 3.7

Calluna- Cladonia P5 251.61 524.73 0.739 3.7

Molinia caerulea 234.71 587.95 -.32 0.626 3.8 P2,3,4,6,7,8

3.8.4 Statistical and uncertainty analysis

Standard errors of the estimates for RTOT, PG and NEE, are shown in brackets unless otherwise stated. Standard deviation is not shown for T5 which was recorded continuously at site rather than plot level. Positive values indicate a loss of carbon from the site to the atmosphere and negative values indicate an uptake of carbon by the site. Statistical uncertainties associated with the models used for annual PG and

RTOT reconstruction were estimated using the standard error of estimation (e.g. Renou-Wilson et al., 2016), where the standard error model (see Eq. 3.9) is determined as a percentage of the mean fluxes which is then related to the annual balance.

2 √∑푛 (퐹표푏푠− 퐹푚표푑) Er = 푖=1 (푛−1)∗푛 (3.9)

where Er is the standard error of the model, Fobs is the sampled flux, Fmod is the modelled flux and n is the total amount of sampled fluxes. The error estimate determined denotes the collective effect of random errors due to statistical uncertainties of the measurements and the scatter of the model results. NEE is not directly modelled, therefore error in the annual NEE estimate was determined following the law of error of propagation as the square root of the sum of the squared standard errors of PG and RTOT.

3.8.2 Reconstruction of CO2 fluxes

The annual balance of fluxes was calculated using the models created in conjunction with the continuous recorded environmental data. PPFD, T5 and T10 were measured every ten minutes and obtained from the weather station (Section 3.5). Daily WT was interpolated from sample day measurements. Daily GAI was calculated for each sample plot using equation 3.3 (Section 3.3.6)

3.4 Results

3.4.1 Environmental variables

In the two years in which the measurements took place (August 2014- September 2015) and the annual balances are modelled (March 2014-February 2015), annual rainfall was greater than the 30 year average (Figure 3.5 a), Year 1 being 16% and year 2 23% wetter. In both years highest rainfall was recorded in the November- February period and the lowest in the summer months, as is typical for a temperate climate. The mean air temperature at Athenry Meteorological Station in 2014 was 9.9°C and in 2015 9.4°C. The highest temperatures were recorded in the June- July period and lowest between December and February. Maximum soil temperatures in both years were reached in July 2014 and 2015 and minimum in the January- February period (Fig 3.5c). Photosynthetic photon flux density values displayed strong diurnal and seasonal variation. In both sample years, PPFD increased steadily in the early months of the year, peaking in June and decreasing steadily towards December (Fig 3.5b). Daily PPFD values were highest between midday and 2pm and

71 zero at night. Annual PPFD was similar in both 2014 and 2015. Water table depth was relatively stable throughout the period of the study. All sample plots experienced a fall in water table during the summer months (Fig 3.6). The largest decrease was observed in P7 a Molinia caerulea dominated plot, in October 2014, where WT fell to -56 cm. The lowest (-27.15, -18.76) mean annual WT was observed in P7 in both 2014 and 2015, while the highest was observed in P3 (-7.2, - 5.71), also a Molinia caerulea dominated sample plot.

400 (a) Year 1

300 Year 2

200 PPT (mm) PPT 100

0 Jan Feb March April May June July Aug Sept Oct Nov Dec

2500 )

-1 (b) 2000 sec -2 1500

1000

500 PPFD (umol m PPFD(umol 0 Mar Jul Nov Mar Jul Nov Mar

25 (c) 20

5 Soil temperature (°C) temperature Soil

0 Mar Jul Nov Mar Jul Nov Mar 2014 2015 2016

Figure 3.5 Climate data for Pollagoona, Co Clare, during the sample period.

(a) monthly average rainfall 2014-2015 (mm) (Met Eireann, Gort rainfall Station),

(b) Photosynthetic photon flux density (PPFD, μmol m-2 s-1), and (c) soil temperature (°C) at 5cm depth from March 2014-March 2016. Dark circles and line indicate 30 year average (1984-2014, www.met.ie). 73

-10

-20

-30

-40 Water table depth (cm) depth table Water -50

-60 Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan Feb Mar Month Figure 3.6 Interpolated water table depths in all study plots from March 2014 until February 2015.

3.4.2 Vegetation dynamics

A strong seasonal progression in GAI was evident in all sample plots throughout the study (Figure 3.7). Plant growth increased throughout the springtime and early summer due to rising soil temperatures and increased PPFD values. Peak GAI occurred in midsummer before decreasing through the autumn due to Molinia caerulea senescence. Throughout the winter months, GAI remained above zero, signifying the presence of evergreen species in all sample plots. The highest variation in GAI occurred in plots with high Molinia caerulea density and least change in GAI observed in the Eriophorum/ Sphagnum plot.

2.5

2.0 ) -2 m 2 1.5

1.0

Green Area Index (m Index Area Green 0.5

0.0

Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan Feb Mar

Month

Figure 3.7 The seasonal development of Green area index in each of the 8 sample plots in Pollagoona.

3.4.3 On-site CO2 fluxes

Measured CO2 fluxes displayed temporal and spatial variation across the study site. Carbon dioxide fluxes followed the seasonal dynamics of soil temperature and water table. Fluxes were low throughout the winter months, increasing steadily until -2 -1 peaking in midsummer. The measured RTOT, ranged between 1055.4 mg CO2 m h -2 -1 -2 - and 64.8 mg CO2 m h while NEE ranged between -2000 and -64.7 mg CO2 m h 1 in the non Molinia caerulea dominated sample plots throughout the study period -2 (see Figure 3.7), however larger fluxes (RTOT between 1734.4 and 41.8 mg CO2 m -1 -2 -1 h , NEE between -2800 and 721.3 mg CO2 m h ) were recorded in the Molinia caerulea plots (Fig 3.9). A significant fluctuation in daily NEE was observed on sampling days. This is most likely caused by changing irradiation levels during the day and variation in GAI between the microsites. Both the Eriophorum vaginatum – Sphagnum and Cladonia portentosa - Calluna vulgaris microsites were C sinks at measurement time while the Molinia plots were both sinks and sources. Highest

75 uptake and losses were seen in the late summer months. Fluxes declined after July in most plots as temperatures fell. The greatest decrease in uptake was seen in the Molinia microsite as the plant senesces and so photosynthesis is greatly reduced. Highest uptake was seen in the Eriophorum vaginatum-Sphagnum microsite in the winter months, reflecting their evergreen growth.

Strong seasonality was also evident in RTOT all microsites (Fig 3.8 and 3.9). Fluxes increased rapidly in spring as soil temperatures increased and vegetation growth began. Greatest losses were in seen in the late summer months with the Molinia -2 -1 microsite recording the highest RTOT (1734.4 mg CO2 m h ).

Inferred PG rates (NEE + RTOT) deisplayed strong seasonality during the study period

(Fig 3.8 and 3.9). In February-March, PG values were close to zero in all microsites but increased rapidly after this time reaching maximum production in late summer and subsequently decling in the autumn. The highest values, around -2800 mg CO2 m-2 h-1 were recorded within the Molinia microsite.

More than one sample was taken per microsite per day. Each point on the graph represents an individual measurement event and so if more than one point accors on the graph on the same day it is not an error.

-2 -1 Figure 3.8 Observed Net Ecosystem Exchange (NEE) (mg CO2 m h ) and -2 -1 ecosystem respiration (RTOT) (mg CO2 m h ) within the Eriophorum vaginatum – Sphagnum and Cladonia portentosa - Calluna vulgaris communities at Pollagoona -2 -1 Co Clare. Gross photosynthesis (PG) (mg CO2 m h ) was calculated as the sum of

NEE and RTOT measurements. Positive values indicate a loss of CO2 to the atmosphere while negative values indicate CO2 uptake to the peatland. Note differences in scale on the y- axis. Fluxes were graphed on Julian Day.

-2 -1 Figure 3.9 Observed Net Ecosystem Exchange (NEE) (mg CO2 m h ) and -2 -1 ecosystem respiration (RTOT) (mg CO2 m h ) within the Molinia caerulea dominated communities at Pollagoona, Co Clare. Fluxes were graphed on Julian Day.

3.4.4 Model performance

The strength of the relationship between RTOT and PG and the environmental variables differed between the vegetation communities. Portrayed graphically below are the responses of bot RTOT (Fig 3.9) and PG (Fig 3.10) to the modelled variables. When modelling, the controlling variable also differed between vegetation communities. Photosynthetic photon flux density alone accounted for 25% of the variability in the Molinia caerulea dominated communities. The addition of soil temperature and GAI (Eq. 3.12) increased the explanatory power of the model to 68%. Eriophorum vaginatum – Sphagnum and Cladonia portentosa - Calluna vulgaris behaved differently to Molinia caerulea dominating plots. Photosynthetic photon flux density accounted for 42-55% of the variation while the addition of soil temperature at 5 cm to the model improved Eriophorum vaginatum – Sphagnum to 79% and further incorporation of GAI to the model raised the explaining power of Cladonia portentosa - Calluna vulgaris to 84%.

2000 2000 1800 1800 1600 1600 ) )

-1 1400 -1 1400 hr hr

-2 1200 -2 1200 m m 2 1000 2 1000 800 800

(mg CO (mg 600 CO (mg 600 TOT TOT

R 400 R 400 200 200 0 0

-50 -40 -30 -20 -10 0 10 2 4 6 8 10 12 14 16 18 20 22 Water table depth (cm) Temperature at 5cm depth (°C)

Figure 3.10 Measured RTOT in response to both temperature at 5cm depth and water table depth.

1000 1000 1000

0 0 0 ) ) ) -1 -1 -1 hr hr hr -2 -1000 -2 -1000 -2 -1000 m m m 2 2 2

-2000 -2000 -2000 (mgCO (mgCO (mgCO G G G P P P -3000 -3000 -3000

-4000 -4000 -4000 0 500 1000 1500 2000 2500 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0 2.2 2 4 6 8 10 12 14 16 18 20 22

Photosynthetic Photon Flux Density Green Area Index Temperature at 5cm depth (°C)

Figure 3. 11 Measured PG in response to PPFD, Green Area Index and temperature at 5cm depth.

At the Eriophorum vaginatum– Sphagnum and Cladonia portentosa- Calluna vulgaris sample plots, soil temperature at 5 cm was the sole explaining variable in the RTOT models (Eq. 3.5) and explained between 63 and 74% of the flux variability. The addition of water table improved the explanatory power of the model for the Molinia caerulea plots from 54 to 63%. The models tended to overestimate small and underestimate high fluxes (Fig 3.12).

Figure 3.12 Relationship between observed and modelled RTOT and PG on Pollagoona

3.4.5 Annual CO2 balance

The annual CO2-C balance for March 2014- February 2015 from the study plots was -2 estimated as an average of the microsite plots NEE as 131.6 ± 298.3 g CO2- C m yr1. The carbon balance in individual sample plots varied spatially across the site. Individual annual balances were averaged to calculate an annual balance. The Eriophorum- Sphagnum microsite was the greatest annual source (184.8 ± 82.8 g -2 -1 CO2- C m yr ) while the Cladonia- Calluna microsite was a modest sink (-142.8 ± -2 -1 91.4 g CO2- C m yr ) for the year studied. The Molinia microsite acted as a source -2 -1 of 168.4 ± 229.5 g CO2- C m yr , showing the great variation between the individual plots in the vegetation type (Table 3.3). Net ecosystem exchange generally followed the water table gradient, annual losses were highest in those collars whose WT dropped below 20 cm depth consistently during the summer -2 -2 months; P7 (243.2 g CO2- C m ), P6 (313.4 g CO2-C m ) and P4 (337.3 g CO2-C m-2 h1). Water table depths deeper than 10 cm were consistently recorded for P5, -2 despite this, it acted as a C sink over the course of this study, (-142.8 g CO2- C m ), suggesting water table is not the controlling variable on this site. Both the Eriophorum- Sphagnum and Molinia caerulea dominated communities acted as small C sinks during the summer months (Fig 3.13). Maximum rate of uptake from -2 -2 the Eriophorum- Sphagnum microsite was 26.1 g CO2-C m , -44.8 g CO2-C m -2 from the Molinia microsite, both in June 2014 and -161.8 g CO2-C m from the Cladonia-Calluna plot in July 2014. Maximum C release from all plots was found in -2 October, 51.2 g CO2-C m from the Eriophrium- Sphagnum microsite, 54.9 g CO2-C -2 -2 m from the Cladonia- Calluna microsite and 41.1 g CO2-C m from the Molinia microsite. A strong seasonal variation was seen in modelled PG, RTOT and NEE. Highest fluxes were seen in all plots in the summer season (May to August) while lowest modelled fluxes appeared in the winter season (November to February) (Fig

3.13). Gross photosynthesis remained below zero, although PG values may have been small, throughout the winter period during daylight hours for the duration of the study indicating that plants were photosynthesising all year round.

(a)

(b)

(c)

2014 2015

Figure 3.13 Average monthly modelled gross photosynthesis (PG), ecosystem -2 -2 respiration (RTOT) and net ecosystem exchange (NEE) (g CO2- C m month ) for (a) Eriophorum- Sphagnum (b) Cladonia- Calluna, and (c) Molinia caerulea sample plots.

Table 3.4 Summary of annual sums and averages of variables from the Eriophorum- Sphagnum (b) Cladonia- Calluna, and (c) Molinia caerulea sample plots. Standard errors of estimates are shown in brackets for all fluxes RTOT, PG and NEE. Positive values indicate a loss of carbon from the site and negative values indicate an uptake of carbon by the site.

T5 WT PG RTOT NEE Eriophorum- 12.6 -10.2 -1089.8(68.1) 1274.7(47.1) 184.8 (82.8) Sphagnum Cladonia- Calluna 12.6 -18.9 -1179.82(86.3) 1036.98(30.1) -142.84(91.4) Molinia 12.6 -13.4 -954.4 (204.7) 1122.9(103.8) 168.4(229.5)

T5 = Soil temperature at 5 cm depth in °C, WT = water table depth in cm. NEE = net ecosystem exchange, PG = gross photosynthesis production, RTOT = ecosystem respiration

3.5 Discussion

3.5.1 Controls of CO2 fluxes

The controlling factors of the NEE components we observed, i.e. soil temperature, WT depth and GAI are well documented (Bubier et al., 1998; Tuittila et al., 2004; Laine et al., 2007a; Pearson et al., 2015). In this study, a positive relationship between temperature, water table and respiration was determined despite the poor fit of the regression model in some cases (Table 3.1 and 3.2). The strong temperature dependence of RTOT has been described in many studies (Silvolia et al., 1996; Bubier et al., 1998; Lafleur et al., 2005). High temperatures stimulate microbial activity emitting greater volumes of CO2 (e.g. Frolking and Crill 1994; Silvola et al. 1996). The chemical structure of substrates and nutrients available also impact on the heterotrophic respiration response to temperature (Updegraff et al. 2001; Blodau et al. 2004). Near surface temperatures displayed a more accurate relationship with

RTOT than deeper soil temperatures, also described by Lafleur et al., (2005) and Minkkinen et al., (2007).

Many previous studies have supported the role of WT in regulating soil respiration (Tuittila et al., 2004; Riutta et al., 2007b) although others have questioned or rejected

83 it (Parmentier et al., 2009; Dimitrov et al., 2010). Water table controls the depth of the oxic layer and therefore the volume of peat in which aerobic decomposition can occur (Baird et al., 2013). Our results suggest that water table is less important than temperature in explaining respiration dynamics in a short term study on a recently rewetted blanket peatland forest with a low water table. The lack of a relationship with the Eriophorum- Sphagnum and Cladonia- Calluna and weak relationship observed between the Molinia plots and WT (Table 3.2) for soil respiration may be due to the depth to which the WT falls in some plots. Once the water table is lowered to a certain depth, sometimes reported as ≈ 30 or 61 cm, soil respiration no longer increases in response (Silvolia et al., 1996; Chimner and Cooper, 2003; Mäkiranta et al., 2009). A stronger relationship would be expected on this site between respiration and WT should the site become wetter and the WT increase above the depth where respiration ceases to respond to WT fluctuations, hopefully lowering respiration from the peat. Another conclusion that can be drawn is that interactions which govern ecosystem processes are complicated and highly site specific.

Vegetation has been suggested as a dominant influence on CO2 sequestration; vegetation cover and successional stage which a site exhibits determining whether it is a sink or source (Samartitani et al., 2011; Urbanova et al., 2012). Vegetation in Pollagoona primarily consists of Molinia caerulea under which some Sphagnum spp. and Eriophorum spp. can be found interspersed with Cladonia and Calluna in the drier areas. Eriophorum vaginatum is influential in C sequestration as it has a high rate of PG throughout the growing season and high respiration rate (Tuittila et al., 2000b; Marinier et al., 2004). Eriophorum cover is low on this site and so the influence of other factors is most likely negating any of its potential effects on C sequestration.Molinia caerulea is the dominant species growing on this rewetted site. It is well adapted to growth in nutrient poor conditions and is known to be a successful and vigorous competitor on drained peatlands (Taylor et al., 2001). Species diversity on this site is likely reduced due to the high production and efficiency of Molinia caerulea (Hájová et al., 2009), limiting the ability of other peat forming species such as Sphagnum spp. to colonise Pollagoona. However, despite this, all sample plots, acted as C sinks during the summer months, providing optimism for the future sink potential of the site. Species diversity must be increased

84 and Molinia dominance reduced in order to encourage the C sink potential. This will rely on the ability of the site to maintain a stable, high water table.

Peat decomposition in temperate climates is supported by mild winters (Höper et al.,

2008), potentially increasing annual CO2 emissions above those in boreal climates. Therefore the need for optimal growth of peat forming plants during the summer is very great. Substantial CO2 emissions from peatlands during winter is a well- documented occurrence (Alm et al., 1999a; Lund et al., 2007; Hendricks et al., 2007), and although Pollagoona is densely covered with vegetation which photosynthesis, primarily through the summer months; not enough C is stored to compensate for winter CO2 losses unlike Sottocornola and Kiely (2005). Temperatures were high and water tables low during the period of the study providing optimal conditions for soil respiration. Water tables were low and although respiration was no longer responding to WT fluctuations due to this and therefore showing a weak relationship, losses through respiration would still be high due to the low WT depths. Peatlands greenhouse gas fluxes are sensitive to weather conditions (Roulet et al., 2007; Urbanova et al., 2013) and so the annual weather patterns will influence the future success of rewetting on this site. Long term studies are needed to track the development of the peatland following restoration and determine its response to changes in weather.

3.5.2 Temporal variation in CO2 exchange

The seasonal variation which we both observed and modelled (Figure 4.2, 4.3, 4.6) is a well recorded trend in temperate conditions (Laine et al., 2007a; Wilson et al., 2007; Wilson et al., 2013). In the study period, fluxes were highest in June, July and August when the greatest GAI and highest PPFD levels are present on the study site

(Fig 3.5b, Fig 3.7, and Fig 3.13). The seasonal changes in PG are mostly controlled by PPFD and GAI rather than temperature which control RTOT (Bubier et al., 1998; Griffis et al., 2000). Fluxes were smallest in the winter months at the coldest period of the year. Soil respiration continued and plant respiration was reduced as some plants senesced in the winter. The presence of evergreen species such as Sphagnum spp. and Eriophorum vaginatum, Eriophorum angustifolium, Campylopus species,

Polytrichum commune and Hypnum jutlandicum on Pollagoona maintained low levels of PG on the site throughout the winter period (Figure 4.6). The contribution of wintertime (November-March) photosynthesis was 3.3% of annual photosynthesis, smaller than respiration which totalled 14.7%. The winter RTOT flux observed here are within the range reported by Laine et al., (2007a) in a temperate natural peatland, lower than found by Alm et al., (1999b) for natural and drained peatland and similar to that of Leppälä et al (2011a) reported in boreal conditions. In the snow free conditions of the temperate climate high winter time respiration levels are anticipated as were observed by Wilson et al., (2007) in a restored Irish peatland, however, Huth et al., (2012), also on a temperate peatland, found wintertime emissions similar to boreal sites due to a harsh cold snap during the study period. Carbon dioxide emissions from the temperate region may drop in response to the colder temperatures experienced during winter (e.g. Beyer and Höper 2015), but unlike Huth et al., (2012), a harsh cold snap was not experienced in this study during the winter period to reduce the CO2 fluxes to the extent of boreal peatlands. Higher WT values than in those recorded in summer may have contributed to reduce peat respiration during winter in this study as levels rise to depths where they influence soil respiration. Plant respiration may also be greatly reduced as much of the vegetation (i.e. Molinia caerulea) dies back in the autumn (Figure 3.7)

3.5.3 Carbon balance of Pollagoona

Long term monitoring of rewetted peatland is required to trace the development of the peatland and successional changes following rewetting. As rewetted peatland forestry has rarely been reported on previously, comparisons will also be made with rewetted industrial peatland or drained peat soils. The results of this study show that previously forested peatland, eight years following rewetting, remains a strong -2 -1 carbon source (131.6 ± 298.3 g CO2-C m yr ). Annual fluxes are similar to those found in other studies (Komulainen et al., 1999; Yli- Petäys et al., 2007) and much -2 -1 lower than those of drained, clear-felled sites; 248-515 g CO2-C m yr , 207-539 g -2 -1 CO2- C m yr (Mäkiranta et al., 2007; Minkkinen et al., 2007). Carbon dioxide emissions on Pollagoona, when compared to drained sites, can be presumed to have

86 fallen over the years since restoration. When compared to forested peatland sites -1 -1 which may store up to 20t CO2 ha yr before the first thinning (Black and Gallagher 2010), potential C losses are great. However, should the sink function return the long term storage will of more benefit than C stored in trees which will be felled ond their timber go into a shrort term use before burning and decomposition, returning C to the atmosphere. While still not acting as a carbon sink, the site demonstrates an improved situation with reduced emissions. While rewetting is utilised to re-establish the C sink function (Höper et al., 2008), should this be unattainable, rewetting on this site has been able to reduce peat emissions which is a positive step in our attempts to reduce climate change. Initially on drained clear- felled peatland sites which have revegetated quickly, the ability of the vegetation to fix carbon cannot compensate for the high RTOT caused by the availability of highly decomposable fresh organic matter (Mäkiranta et al., 2012), similar to the situation found on Pollagoona where brash material still remains. High CO2 emissions could also be contributed to by the ‘priming effect’ i.e. the decomposition of peat beneath the brash could be increased by the introduction of fresh organic material from the brash. Studies have previously shown (Fontaine et al., 2004; Fontaine et al., 2007), the presence and significance of priming in organic matter decomposition. The influence of the priming effect will decrease as the volume of brash material decreases through decomposition.

Time since rewetting has emerged as an influencing factor in the re- establishment of the C sink function in the peatland. High CO2 emissions immediately following rewetting are not limited to rewetted peatland forest. Rewetted industrial harvested peatland has been shown to produce high levels of CO2 in in the early stages post rewetting (478 and 468 g C m−2 yr-1) (Petrone et al., 2003). Other industrial rewetted sites continue to perform as CO2 sources thirty years post rewetting (Petrone et al., 2003; Samartitani et al., 2011), despite maintaining high water tables throughout the study. The C losses observed were attributed to vegetation characteristics and the ability of the rewetted sites to resist temperature and WT changes.

Older rewetted peat extraction sites have been found to be C sinks (Samartitani et al., 2011). It can be suggested that the C function in this site is similarly time dependent and may yet return. Rewetting is a long term strategy to reinstate the carbon sink function of a peatland. It is dependent on a high water table and the

87 vegetation present on the site (Drösler et al., 2008). While many peat forming plants are present on this site, the mean water table drop at dry periods maintains an aerobic environment in the peat, allowing for peat decomposition and therefore increased

CO2 emissions.

While most of the plots were sources in the year modelled, all vegetation communities acted as sinks in summer months. The greatest sink was the Cladonia- Calluna community followed by the Molinia caerulea dominated plots. Calluna dominated communities have been found to be net sinks in a previous study (Larson et al., 2007). Although Molina colonization and dominace is not an ideal scenario and increased growth of other peatland species such as Sphagnum should be encouraged it has been found in previous work that Molinia caerulea acts as a strong C sink (Urbanová et al., 2012) which may be promising for the restoration potential of the site. Respiration showed great variation throughout the year and between microsites. Gross photosynthesis was high all microsites in the mid-summer months. Photosynthesis levels dropped considerably throughout the autumn and winter as Molinia caerulea senesced (Taylor et al., 2001) and green area was reduced. Net ecosystem exchange remained positive at all times during the winter period similar to other rewetted peatlands in Ireland (Wilson et al., 2007; Renou-Wilson et al., 2016). Being less than ten years old, plants are near the start of the building stage of their life cycle (Gimingham, 1960) and the increasing plant biomass indicates high photosynthesis levels. Height canopies, an influencing factor in C loss from Calluna communities (Dixon et al., 2015) also remained low in this sample plot limiting the

CO2 loss. As Calluna vulgaris height increases photosynthesis per unit respiration increases. Water table levels at the Cladonia- Calluna community remained close to -20 cm for much of the study, inhibiting soil respiration. Rewetted sites which have regained their C sink function maintain a high water table all year (e.g. Tuittila et al., 1999; Wilson et al., 2013) a feature which is missing from this site. Therefore, in order for this site to store C, the water table must be raised and maintained at a higher water table.

Not all C losses have been accounted for in this study. Besides gaseous losses, C is also lost through waterborne carbon fluxes such as dissolved organic carbon (DOC), particulate organic carbon (POC), dissolved inorganic carbon (DIC) and gaseous

CO2 and CH4 , in run-off from the peatland (Evans et al., 2015). Surface waters from

88 peatland catchments, supersaturated with CO2 and CH4 represent a significant pathway connecting C store with the atmosphere (Dinsmore et al., 2009a). Waterborne carbon can contribute greatly to the C balance of a peatland, turning it from a calculated sink or neutral peatland to a carbon source (Roulet et al., 2007).

Dissolved organic carbon, similarly to CO2, originates in the aerobic area above the water table. A significant relationship has been determined between vegetation type and DOC concentration levels (Armstrong et al., 2012), highlighting the influence vegetation has on both gaseous and fluvial C losses. In future work, to accurately account for all C losses, waterborne C analysis should be included.

3.6 Conclusions

Our study highlights that peatland restoration is be a long term process in rewetted forestry greatly influenced by vegetation and the presence of fresh organic matter left on the site following clear-felling. Based on the results of this study, we can conclude that rewetting, while not an instant solution to CO2 emissions from drained peatland reduces C losses considerably and rewetted areas will in time potentially return to a C sink. In order to further reduce peat decomposition, it can be suggested that the water table should be raised further. Winter time respiration fluxes remain low on temperate rewetted peatland sites, despite the mild winters. It can be recommended that in rewetted forestry projects, clear-felled trees should be removed to reduce the volume of fresh organic material on the site encouraging a faster return of the C sink function. Further study should also be conducted on rewetted forestry on peatland sites sometime after rewetting to monitor the development of the site as it returns to a C sink.

Acknowledgements This project is funded by the Environmental Protection Agency (2012-CCRP- PhD. 2). Grateful thanks to Coillte for access to their LIFE 02 and LIFE 09 sites. Thanks to Michael Kenny (Carlow Institute of Technology), Michael Kenna (Noone Engineering, Rathangan, Co Kildare) and Ray Byrne for assistance in equipment construction and David Wilson and Flo Renou Wilson for assistance with field measurements and analysis of results.

Chapter Four

Carbon dioxide emissions in a raised bog after clearfelling and rewetting

4.1 Abstract

Natural peatlands are important sinks of carbon (C) and vital in the global C cycle. Drainage for forestry converts peatlands from sinks to sources of carbon dioxide

(CO2). Forest drained peatlands in Ireland have been rewetted with the aim of returning these ecosystems to their natural condition. Rewetting is utilized to return conditions for C sequestration to the peatland and has been recognised as an important climate change mitigation tool. Since rewetting, water table levels have risen on the study site and more typical peatland vegetation species such as Sphagnum capillifolium, Sphagnum papillosum, Eriophorum vaginatum, Eriophorum angustifolium Calluna vulgaris, Erica tetralix and Hypnum jutlandicum have colonised the site. The aim of this study was to investigate the controls on gross photosynthesis (PG), ecosystem respiration (RTOT) and net ecosystem CO2 exchange (NEE) in a rewetted raised peatland forest and to determine if the C sink function of the peatland had been restored three years after rewetting. Carbon dioxide fluxes were measured using the chamber method. Four microsites comprising the dominant vegetation on the study site were established and measured for one year. All microsites were annual sources of C. Temporal variation was most obvious in PG which decreased substantially in the winter months during which leaf area and photosynthetic photon flux density is at its lowest. The annual CO2–C balance was 585.3± 241.5 g m-2 year-1. Site topography and the large volume of brash remaining on site were influential in determining the annual C balance with the greatest losses seen from those sites in the plough furrows and brash areas and lowest emissions found on the ridges. Our results show that the presence of a dense fresh organic matter layer following rewetting maintains high carbon losses from forestry sites delaying their return to C sink.

4.2 Introduction

Peatlands cover just 3% of the world’s land surface; despite this they store as much C as all terrestrial biomass and double that of global forest biomass (Parish et al., 2008). Primarily due to drainage for agriculture and forestry as well as peat mining for fuel and horticulture, an estimated 80 million ha worldwide has been lost (Joosten and Clarke, 2002). Peatlands develop due to an accumulation of organic matter deriving from decaying plant and animal material caused by impeded decomposition under saturated, anaerobic conditions (Moore, 1987; Renou-Wilson et al., 2011). Peatlands are vulnerable to climatic and hydrological changes such as drainage, which can convert them from atmospheric C sinks to sources of CO2 (Silvola et al., 1996; Strack, 2008; Ojanen et al., 2010). Peatlands in Ireland represent Europe’s sixth largest peatland area (Montanarella et al., 2006), covering approximately 20% of the country (Connolly and Holden, 2009). Despite the abundance of peatland, only approximately 10% of the original raised peatland is deemed to be in a near natural condition (NPWS, 2013).

Raised peatland is formed in areas with impermeable soil such as hollows or lake basins and a consequently high water table (Gardiner and Radford, 1980); occurring frequently in the Irish midlands (Hammond, 1981). Afforested peatland in Ireland is estimated at 322 000 ha, over 10% of which is on raised peatland (NFI, 2013). Forest drainage causes a detrimental shift in ecosystem characteristics such as aeration of the surface peat, nutrient conditions, soil bulk density and peat subsidence as well as in microbial activity (Minkkinen and Laine et al., 1998; Laine et al., 2006) and triggers the replacement of peatland vegetation by woodland species (Korpela 1999; Laine and Vanha- Majamaa 1992).

Peatlands occupy an important position in mitigating climate change owing to their current C store and future sink potential (Yu et al., 2011). In their natural condition, peatlands are recognised for their high conservation value (Charman 2002), and many nations are now making efforts to restore degraded peatlands (Joosten et al., 2012). Restoration attempts are common in North America (Rochefort et al., 2003) and in Europe (Hendricks et al., 2007; Koskinen et al., 2016; Wilson et al., 2016b). Restoration practises range from vegetation management (Rochefort et al., 1997;

Anderson et al., 2009) to rewetting (Wilson et al., 2007b; Zhang et al., 2012). Rewetting is defined here as ‘the deliberate action of raising the water table on drained soils to re-establish water saturated conditions’ (IPCC, 2014b). Rewetting initially aims to limit further peat degradation (Schuman and Joosten, 2008), followed by re-establishing a peatland ecosystem similar to the one which was degraded (Konvalinková and Prach, 2014), primarily through drain-blocking to raise the water table (Bragg, 2011). Conditions at rewetting initiation; vegetation, hydrological functioning, past management, pH and length of drainage (Höper et al., 2008; Rydin and Jeglum 2013), impact on the future trajectory of a peatland returning to a C sink. Once rewetted, GHG dynamics are affected by a number of factors; time since rewetting, pH, nutrient status, peat type, vegetation present and hydrology, which suggests an individual approach is needed for each rewetting project (Wilson et al., 2016a) such as in drain block selection, berm building or vegetation reintroduction. Evidence so far has proven rewetted peatlands react unpredictably to rewetting in terms of GHG dynamics and results have shown little consistency. Some studies report the rapid return of the C sink function (Wilson et al., 2011; Urbanová et al., 2013); while in other instances high CO2 emissions remain for years post rewetting (Wilson et al., 2007b; Samartitani et al., 2011). Water table dynamics have been found to behave differently to natural peatland even many years post rewetting (Holden et al., 2011), presenting challenges for peatland restoration. Work by Komulainen et al., (1999) in Finland has found that soon after restoration, vegetation succession was initiated and

CO2-C balance was developing towards that of natural peatland. Following previous work on rewetted peatland on both different land uses and world regions this study aims to: (a) identify the controlling variables of CO2 fluxes on such rewetted peatland forestry sites, (b) estimate the annual CO2 balances for a range of rewetted microsites at a newly rewetted raised peatland forest in the temperate region and (c) to compare our findings with other rewetted peatland sites.

4.3 Materials and Methods

4.3.1 Site Description Scohaboy Bog is a raised bog located 4 km south-east of Borrisokane in Co. Tipperary (52°59' N, 8°02' W) and lies at an altitude of 78 m above sea level (Fig 4.1). The nearest synoptic weather station to Scohaboy is located at Gurteen College, 7 km north of the site. The climate is temperate and humid and the mean annual air temperature over the last 30 years was 9.8 °C and mean annual rainfall was 948.2 mm (Met Éireann, 1981- 2010, average at Birr and Gurteen Meteorological Stations, http://www.met.ie/climate/monthly-data.asp?Num=1875).

Figure 4.1 Location of the Scohaboy Bog site.

Figure 4.2 Site map of Scohaboy rewetted area (outlined in red). Drain blocking was also conducted on the unplanted bog

The site includes a large natural bog (approximately 52 ha), cutover and previously forested peatland. The restoration site which was mainly planted with Sitka spruce (Picea sitchensis [Bong] Carr) in the 1980s, totals 71.80 ha. Agricultural land borders the site on all boundaries, except for the presence of coniferous forestry to the north. Drains were dug and an area of natural bog ploughed, in efforts to create conditions suitable for forestry. In brief, restoration activities conducted were clear felling to waste of the trees on approximately 19 ha in 2011, windrowing the trees to limit the area of bog covered by trees thereby allowing for maximum recovery of bog vegetation and blocking of drains with plastic piling on both the clear-felled peatland and the remaining natural bog and installing peat dams on the deforested area in 2013 (Coillte, private correspondence). Brash covered approximately 40% of the site. Furrows and ridges are still evident on the rewetted bog, creating contrasting environments for vegetation colonisation. The furrows remain permanently flooded while dryer areas occur on the ridges between furrows and drains. The rewetted former forested area is separated from the natural bog by a fire break area consisting of wet pools while there are 2.5 ha of riparian woodland to the north that has been

98 retained and acts as a lag zone, being at the transition between peat and mineral soil and including a stream.

Scohaboy vegetation was consistent with that found predominately in raised peatlands, as described by Goodwin and Conway (1939). Plots were labelled S1- S8; S1- S4 were located in the wettest area of the site and S5- S8 were located in a drier area. Sample plots (shown in Fig 4.3) in the furrows (S1, S2 and S5) were the wettest and were covered by a dense carpet of Sphagnum capillifolium and Sphagnum papillosum, along with Eriophrium vaginatum. Erica tetralix also grew in the wetter sample areas. In sample plots where the water table remained close to the surface for most of the sample period, Eriophrium vaginatum grew abundantly. Raised plots situated along the tree lines (S3, S6) were drier throughout the sample period. The vegetation growing in these plots, i.e. Calluna vulgaris and Cladonia species; reflect these drier conditions. Sample plots situated in the brash line (S4, S7 and S8) provided conditions quite unlike the other sample plots. Little or no vegetation was present at the start of the study, however over time Sphagnum papillosum and Hypnum jutlandicum started to colonize the plots.

Table 4.1 Vegetation species recorded in the study plots. Species are listed in descending order of dominance.

Eriophorum – Eriophorum vaginatum Cladonia portentosa- Brash Sphagnum (S2) Mosses (S3+ S6) (S4, S7 +S8) (S1 +S5) Sphagnum capillifolium Eriophorum vaginatum Cladonia portentosa Hypnum jutlandicum Sphagnum papillosum Hypnum jutlandicum Calluna vulgaris Sphagnum papillosum Eriophorum vaginatum Sphagnum capillifolium Hypnum jutlandicum Erica tetralix Cardamine pratensis Cardamine pratensis ssp. Paludosa ssp. Paludosa Campylopus spp. Eriophorum vaginatum Eriophorum angustifolium Sphagnum papillosum Sphagnum capillifolium

Figure 4.3 Scohaboy microsites in May 2014. Top from left to right, S1 to S4, bottom from left to right, S5 to S8.

4.3.2 Study Site

The site was established in December 2013, three months prior to measurements commencing. Eight steel collars (60 cm × 60 cm × 30 cm) were inserted into the peat in two parallel transects collars, 60 meters apart, 4 collars in each, running from East to West. This design was used to capture the observed water table variation (observed by water table sampling and vegetation present) and vegetation variation between the furrows, ridges and brash lanes. The collars were topped with a 3cm deep channel that was filled with water to provide an airtight seal during measurements. Boardwalks were constructed beside each collar to minimise disturbance and damage to peat surface and plant cover during measurements.

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4.3.3 CO2 flux measurements

Light and dark chambers were used to measure instantaneous net ecosystem exchange (NEE). Net ecosystem exchange was measured using a static polycarbonate chamber (60 cm × 60 cm × 30 cm) fitted with a battery operated fan, whose function was to mix the air within the chamber head-space. The chamber was connected to a cooling system which maintained the internal temperature to within 1 ºC of the ambient temperature. Carbon dioxide fluxes were measured using a portable infrared gas analyser (EGM- 4, PP Systems, UK). In order to measure instantaneous NEE, CO2 concentration initially was measured at 15 s intervals over a period of 60 s to 180 s under ambient illumination. The chamber was vented for a short time following the initial measuring of CO2 in each collar. To better establish a relationship between PPFD (photosynthetic photon light density) and photosynthesis, artificial shades were used to obtain fluxes under a greater range of PPFD. Following light measurements, the chamber was then replaced on the collar and covered with an opaque material and CO2 fluxes measured to determine total ecosystem (heterotrophic and autotrophic) respiration (RTOT) (method as per Wilson et al 2007, Tuittila et al., 1999 and Tuittila et al., 2004). Carbon dioxide flux rates were calculated from the linear change in gas concentration as a function of time. The ecological sign convention, in which fluxes from the atmosphere to the biosphere are negative, was used. Gross photosynthesis

(PG) was estimated as the sum of flux rate values measured in light (NEE) and dark

(RTOT). Carbon flux measurements were conducted on a fortnightly basis except for one month period in December 2014- January 2015, from August 2014 until September 2015. Measurements were made between 8.00 am and 5.00 p.m. in a random pattern. Two measurement rounds were conducted on most sampling days.

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PPFD sensor

Raising Collar

Cooling system

Thermometer

EGM- 4

Figure 4.4 CO2 flux chamber and raising collar, equipped with a cooling system and fan. An infrared gas analyser (EGM- 4) measures CO2 concentration. PPFD and air temperature are measured inside the chamber.

4.3.4 Environmental Variables

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HYDROKIT Dorset. The water table data was interpolated between measurements in order to describe water table patterns across the seasons. A meteorological station (Watchdog 1400 Micro station) was installed on each site measuring PAR and soil temperature at 5 cm and 10 cm at ten minute intervals throughout the duration of the study.

4.3.6 Green Area Index

Green area index (GAI), which represents the green plant material, was used in gas flux modelling to relate gas fluxes to the seasonal changes in plant growth and senescence. The GAI of each sample plot was calculated based on Wilson, et al., 2007. Each sample plot was placed in vegetation communities’ representative of those present across the restored peatland. Within each site, five sub sample plots (8cm × 8cm) were established, one plot in each corner of the collar and one in the middle of the collar. Therefore the vegetation range in the microsite was captured. At fortnightly intervals, the total number of leaves and stems of individual plant species within the subsample plots were counted and used to estimate the average number of leaves and stems of each species per m2. In order to avoid disturbance to the plants inside the sample plots, 3 individual plants of each species outside the sample plots, were tagged and their leaves measured. These tagged plants were similar in age and stature to those inside the sample plots. The length and width of each leaf was measured and used to estimate leaf area. The area of the leaves was estimated using species-specific formulae, based on the geometric shape of the leaf (e.g. ellipse, circle, rectangle). These measurements were made on the same day as leaves were counted. If present, the green surface area of the plant stem was calculated using the formula for surface area of a cylinder; Green stem area= (2*π*r)*h (Eq. 3.1) where r is the radius and h is the plant stem height. A percentage vegetation cover was also conducted at the same time on all subplots. The GAI of plants with irregular growth patterns or irregular leaf edges was estimated from this percentage cover.

The average stem area (SA) and leaf area (LA) were calculated to give a representative leaf and stem size for each species on each sample day throughout the study period.

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The green area of each species, i (m2 m-2) was calculated using the following formula;

GAi = (LA * Ln) + (SA * Sn) (Eq. 3.2) 2 where LA is the average leaf area of the three tagged plants (m ), Ln is the average 2 2 leaf area per m , SA is the average stem area of the tagged plants (m ) and Sn is the average number of stems per m2.

4.3.7 Modelling the annual development of leaf area index

푗푢푙푖푎푛−푥 (0.5( 푚푎푥)2) Daily GAI: y+ GAImax e 푏 , (Eq. 3.3)

4.3.8 Data Analysis

4.3.9 Reconstruction of CO2 fluxes

104 was interpolated from sample day measurements. Daily GAI was calculated for each sample plot using equation 3.3 (Section 3.3.6)

4.3.10 Modelling of CO2 Exchange

(Stryer, 1988). Non- linear regression has been widely used in modelling PG, owing to the use of the hyperbolic relationship (Alm et al., 2007; Tuittila et al., 2004;

Wilson et al., 2007; Urbanová et al., 2012). PG usually has distinctive responses to the remaining measured variables: GAI- linear, temperature- exponential and WT-

GAI was not used in RTOT modelling as no relationship was determined. Nonlinear regression models were initially constructed separately for each individual collar. Plots were grouped into microsites based on vegetation characteristics, WT levels and absence or presence of brash. SPSS 22 for Windows statistical package (SPSS, Inc.) was used in flux modelling. The models were parameterized for each vegetation community as each community differed in response to environmental variables. Gross photosynthesis is very dependent on PPFD and is regularly described using the Michaelis-Menten relationship. Water table and PPFD were used as explanatory variables for the PG

105 models. Water table was added to the model as a Gaussian term (Eq. 4.1). The model was adapted from Wilson et al. (2013). Only variables which increased the explanatory power of the model were accepted. Neither temperature nor GAI improved the models and so were excluded. Model acceptance was based on the following criteria; (a) statistically significant model parameters (p<0.005), (b) highest possible coefficient determination and (c) lowest possible parameter error. Gross photosynthesis was not modelled for brash collars.

The PG model used was as follows;

푃푃퐹퐷 푊푇−푊푇표푝푡 2 PG= Pmax ( ) ∗ (퐸푋푃(−0.5 (( ) ) (Eq. 4.1) 푃푃퐹퐷+푘푃푃퐹퐷 푊푇푡표푙 where, PG is gross photosynthesis, Pmax is maximum photosynthesis, PPFD is photosynthetic photon flux density, kPPFD is the value at which PG reaches half its maximum, WT is interpolated water table depth and WTopt is the optimum water level and WTtol is the water level tolerance for photosynthesis. The PG model parameters for each microsite are given in Table 4.2.

2 Table 4.2 Estimated parameter values, goodness of fit (r ) for PG. Number of the equation used is given for each sample plot. Plots are labelled S1- S8.

Sample Plot PG

2 Pmax kPPFD WTopt WTtol R Eq. Eriophorum- Sphagnum S1/5 3274.46 466.92 -12.10 9.68 0.77 4.1

Eriophorum S2 1563.49 194.92 -16.11 9.34 0.64 4.1

Cladonia- Mosses S3/6 3082.72 432.78 -13.59 10.76 0.68 4.1

Ecosystem respiration (RTOT) is strongly influenced by temperature and water table

(Laine et al., 2009b; Juszczak et al., 2013). RTOT was linked to soil temperature at 5 cm (Kelvin) using the exponential relationship described by Lloyd and Taylor (1994)

(Eq. 4.2). In this study, water table displayed a linear relationship with RTOT in some

106 plots, improving the explaining power of the model. GAI did not improve the models and so was excluded. Therefore the RTOT models used for Scohaboy were

1 1 [퐴 ∗ 퐸푋푃[C ( − )] (Eq. 4.2) 푇푟푒푓−푇0 푇− 푇0

1 1 [퐴 + (퐵 ∗ 푊푇) ∗ 퐸푋푃[C ( − )], (Eq. 4.3) 푇푟푒푓−푇0 푇− 푇0

1 1 [퐴 ∗ 퐸푋푃[C ( − )] + (B*WT) (Eq. 4.4) 푇푟푒푓−푇0 푇− 푇0 where Tref was set at 283.15 and T0was set at 227.13, according to Lloyd and Taylor (1994). The temperature minimum at which respiration reaches zero is denoted by

T0, WT is interpolated water table depth, A, B and C are fitted model parameters.

Soil temperature (T) is given in degrees Kelvin. The RTOT model parameters for each sample plot are given in Table 4.3

2 Table 4.3 Estimated parameter values, goodness of fit (r ), for RTOT. Number of the equation used is given for each sample plot. Plots are labelled S1- S8.

Sample Plot RTOT A B C R2 Eq. Eriophorum- Sphagnum S1,5 460.19 -52.75 212.72 0.449 4.4

Eriophorum S2 367.67 401.60 0.61 4.2

Cladonia- Mosses S3,6 150.92 -6.98 354.26 0.77 4.3

Brash collars S4,7,8 183.43 -11.55 271.41 0.450 4.4

4.3.11 Statistical and uncertainty analysis

Standard errors of the estimates for RTOT, PG, NEE, CH4 and N2O are shown in brackets unless otherwise stated. Standard deviation is not shown for T5 which was recorded continuously at site rather than plot level. Positive values indicate a loss of carbon from the site to the atmosphere and negative values indicate an uptake of

107 carbon by the site. Statistical uncertainties associated with the models used for annual PG and RTOT reconstruction were estimated using the standard error of estimation (e.g. Renou-Wilson et al., 2016), where the standard error model (see Eq. 3.9) is determined as a percentage of the mean fluxes which is then related to the annual balance.

2 √∑푛 (퐹표푏푠− 퐹푚표푑) Er = 푖=1 (푛−1)∗푛 (3.9) where Er is the standard error of the model, Fobs is the sampled flux, Fmod is the modelled flux and n is the total amount of sampled fluxes. The error estimate determined denotes the collective effect of random errors due to statistical uncertainties of the measurements and the scatter of the model results. NEE is not directly modelled, therefore error in the annual NEE estimate was determined following the law of error of propagation as the square root of the sum of the squared standard errors of PG and RTOT.

4.4 Result

4.4.1 Environmental Variables

In the two years in which the measurements took place (August 2014- September 2015) and the annual balances are modelled (March 2014-February 2015), annual rainfall was similar to the 30 year average (Fig 4.5 a), precipitation in year 1 was 4% less and year 2, 19% wetter than the average. The highest rainfall in both years occurred in the winter months and the lowest in June- July, as is typical for this climate. In year 1, mean air temperature at Gurteen Meteorological Station, was 9.9°C and in year 2, 9.5 °C. Maximum temperatures in both years were recorded in July and minimum temperatures in February. Soil temperatures in both years reached their maximum in July 2014 and 2015 and minimum in the January- February period (Fig 4.5c). Photosynthetic photon flux density values displayed strong diurnal and seasonal variation. In both sample years, PPFD increased steadily in the early months of the year, peaking in June and decreasing steadily towards December. Daily PPFD values were highest between midday and 2 pm and zero at night. Annual PPFD was similar in both 2014 and 2015 (Fig 4.5b). Water table depth was relatively stable throughout the period of the study (Fig 4.6). All sample plots

108 experienced a fall in water table during the summer months, although WT remained above -30 cm throughout the sample period. Both the lowest (-19.6, -22.5) and highest (0.37, 1.8) mean annual WT was observed in brash plots in both 2014 and 2015.

250 (a) 200 Year 1

150 Year 2

100 PPT(mm)

0 March April May June July Aug Sept Oct Nov Dec Jan Feb

2500 ) -1 2000 (b) sec -2 1500

1000

500 PPFD (umol m PPFD(umol 0 Mar May Jul Sep Nov Jan Mar May Jul Sep Nov Jan Mar

25 (c)

5 Soil temperature (°C)

0 Apr Jun Aug Oct Dec Feb Apr Jun Aug Oct Dec Feb

2014 2015 2016

Figure 4.5 Climate data for Scohaboy Bog, Co Tipperary during the sample period, March 2014-March 2016. (a) monthly rainfall (mm) (Met Eireann, Gurteen Station), (b) Photosynthetic photon flux density (PPFD, μmol m-2 s-1), (c) soil temperature

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(°C) at 5 cm depth. Dark circles and line indicate 30 year average (1981-2010, www.met.ie).

-2

-4

-6

-8

-10 Water table (cm) table Water -12

-14

-16 Apr Jun Aug Oct Dec Feb Month Figure 4.6 Mean sample day water table in brash plots, Sphagnum/ Eriophorum plots and Cladonia/ Mosses plots, March 2014- February 2015.

4.4.2 Measured CO2 fluxes

Measured CO2 fluxes displayed temporal and spatial variation across the study site. -2 -1 -2 -1 The measured RTOT, ranged between 849.2 mmg CO2 h and 33.7 mg CO2 m h in brash plots, in the Cladonia-Mosses microsite RTOT, ranged between 810 mg CO2 -2 -1 -2 -1 m h and 39.5 mg CO2 m h while NEE ranged between -1137.6 and 127.9 mg -2 -1 -2 CO2 m h , in the Eriophorum microsite RTOT, ranged between 1378.4 mg CO2 m -1 -2 -1 h and 81.3 mg CO2 m h while NEE ranged between -1398.1 and 538.1 mg CO2 -2 -1 -2 -1 m h , however larger fluxes (RTOT between 2525.2 mg CO2 m h and 33.4 mg -2 -1 -2 -1 CO2 m h and NEE between -1165.2 and 1105.9 mg CO2 m h ) were recorded in the Eriophorum- Sphagnum plots located in the flooded plough furrows (Fig 4.7 and Fig 4.8).

A significant fluctuation in daily NEE was observed on sampling days. This is most likely caused by changing irradiation levels during the day and variation in GAI between the microsites. The Eriophorum vaginatum – Sphagnum, Cladonia-Mosses and Eriophorum microsites were C sinks at measurement time while the Brash plots were both sinks and sources. Highest uptake and losses were seen in the late summer

110 months. Fluxes were low throughout the winter months, increasing steadily until peaking near day 200, in all sample plots excluding the Eriophorum plot where the majority of the highest fluxes were recorded after day 250.Fluxes declined after July in most plots as temperatures fell.

Strong seasonality was also evident in RTOT all microsites (Fig 4.7 and 4.8). Fluxes increased rapidly in spring as soil temperatures increased and vegetation growth began. Greatest losses were in seen in the late summer months with the Eriophorum- -2 -1 Sphagnum microsite recording the highest RTOT (2525.2 mg CO2 m h ).

Inferred PG rates (NEE + RTOT) deisplayed strong seasonality during the study period

(Fig 4.7 and 4.8). In February-March, PG values were close to zero in all microsites but increased rapidly after this time reaching maximum production in late summer and subsequently decling in the autumn. The highest values, around -2600 mg CO2 m-2 h-1 were recorded within the Eriophorum microsite.

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-2 -1 Figure 4.7 Observed net ecosystem exchange (NEE) (mg CO2 m h ) and -2 -1 ecosystem respiration (RTOT) (mg CO2 m h ) within the Eriophorum vaginatum – Sphagnum and Cladonia portentosa- Mosses communities at Scohaboy Bog, Co. -2 -1 Tipperary. Gross photosynthesis (PG) (mg CO2 m h ) was calculated as the sum of

NEE and RTOT measurements. Positive values indicate a loss of CO2 to the atmosphere while negative values indicate CO2 uptake to the peatland. Note differences in scale on the y- axis.

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-2 -1 Figure 4.8 Observed net ecosystem exchange (NEE) (mg CO2 m h ) and -2 -1 ecosystem respiration (RTOT) (mg CO2 m h ) within the Eriophorum vaginatum and

Brash communities at Scohaboy Bog, Co. Tipperary. Gross photosynthesis (PG) (mg -2 -1 CO2 m h ) was calculated as the sum of NEE and RTOT measurements. Positive values indicate a loss of CO2 to the atmosphere while negative values indicate CO2 uptake to the peatland. Note differences in scale on the y- axis.

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4.4.3 Model performance

The strength of the relationship between CO2 and the environmental variables varied between the microsites. Portrayed graphically below are the responses of bot RTOT

(Fig 4.9) and PG (Fig 4.10) to the modelled variables. In modelling PG fluxes, the controlling variable differed between vegetation communities. Photosynthetic photon flux density alone accounted for 45%-53% of the variability across all vegetation communities. The addition of water table (Eq. 5.1) increased the explanatory power of the model to 77% (S1 and S5), 64% (S3 and S6) and 68% (S2). No significant relationship between GAI and CO2 flux was observed and so was not included. At

S2, soil temperature at 5 cm was the sole explaining variable in the RTOT models (Eq. 5.2) and explained 61% of the flux variability. The addition of water table to the model (Eq. 54) increased the variability explained to 44.9% (S1 and S5) and 45% (brash collars). Water table was the controlling variable in the Calluna- Mosses sample plots (Eq. 5.3) and the addition of soil temperature at 5 cm improved the explanatory power of the model explaining 77% of the variability.

2500 2500

2000 2000 ) ) -1 -1 h h -2 -2 1500 1500 m m 2 2

1000 1000 (mg CO (mg CO (mg TOT TOT

R 500 R 500

0 0 -30 -20 -10 0 10 20 0 5 10 15 20 25 Water Table depth (cm) Temperature at 5cm depth (°C)

Figure 4.9 Measured RTOT response to both water table depth and temperature at 5cm depth.

114

3000 3000

2500 2500 ) ) -1 -1 2000 2000 hr hr -2 -2 m m 2 1500 2 1500

1000 1000 (mg CO (mg CO (mg G G P P 500 500

0 0

-30 -20 -10 0 10 20 0 500 1000 1500 2000 2500 Water Table Depth (cm) Photosynthetic Photon Flux Density

Figure 140 Measured PG in response to water table depth and PPFD.

The agreement between the modelled and predicted fluxes contrasted across the vegetation communities. The models tended to overestimate small and underestimate high fluxes.

2500 2500 ) ) -1 -1 h h

-2 2000

-2 2000 m 2 m 2 1500 1500 (mg CO (mg

1000 CO (mg G TOT 1000

500 500 Predicted P Predicted

Predicted R Predicted 0 0

0 500 1000 1500 2000 2500 3000 0 500 1000 1500 2000 2500 3000 -2 -1 -2 -1 Measured P (mg CO m h ) Measured RTOT(mg CO2 m h ) G 2

Figure 4.11 Relationship between observed and modelled RTOT and PG on Scohaboy Bog

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4.4.4 Annual CO2 balance

The mean annual C balance across the study plots from March 2014-February 2015 -2 -1 was estimated as 585.3±241.52 g CO2- C m yr . The C balance varied spatially across the site. All sample plots were a source of CO2 during the sample period. In the brash plots, NEE followed the WT gradient, the annual balance was highest at S7 -2 -2 (761.3 g CO2-C m ) and S8 (1136.8 g CO2-C m ), both plots recorded the lowest -2 water tables. Despite a high WT ( -7, -8, 0, -3 respectively), S1 (785.9 g CO2-C m ) -2 -2 -2 S2 (382.3 g CO2-C m ), S4 (559.8 g CO2-C m ) and S5 (668.3 g CO2-C m ) emitted high levels of CO2 while lowest annual emissions occurred at the dryer plots -2 -2 of S3 (25.1 g CO2-C m ) and S5 (363.1 g CO2-C m ). A strong seasonal variation was also seen in modelled PG, RTOT and NEE. Highest fluxes were observed in all plots in the summer season (May to August) while lowest modelled fluxes appeared in the winter season (November to February). Gross photosynthesis remained below zero throughout the winter period during daylight hours for the duration of the study.

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400 200 (b) 300 (a) 100 -1 -1 200

month month 0 100 -2 -2

0 -C m -C m -C -100 2 2

-100 g COg COg -200 -200

-300 -300 3 4 5 6 7 8 9 10 11 12 1 2 3 4 5 6 7 8 9 10 11 12 1 2 300 160

-1 120 100 100 month month -2 -2 0 80 -C m -C -C m -C 2

2 60 -100

g COg 40 g COg -200 20

-300 0 3 4 5 6 7 8 9 10 11 12 1 2 3 4 5 6 7 8 9 10 11 12 1 2

2014 2015 2014 2015 2015

Figure 4.12 Average monthly modelled gross photosynthesis (PG), ecosystem -2 -2 respiration (RTOT) and net ecosystem exchange (NEE) (g CO2- C m month ) for (a)Eriophorum- Sphagnum (b) Cladonia- Mosses, (c) Eriophorum and (d) brash sample plots. Note the different scales on the y axis.

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Table 4.4 Summary of annual sums and averages of variables from Eriophorum- Sphagnum, Cladonia- Mosses, Eriophorum and brash sample plots. Standard errors of estimates are shown in brackets for all fluxes RTOT, PG and NEE. Positive values indicate a loss of carbon from the site and negative values indicate an uptake of carbon by the site.

T5 WT PG RTOT NEE Eriophorum- 13.5 -3.8 -1159.6 1886.7 727.1 Sphagnum (179.9) (240.4) (300.3) Eriophorum 13.5 -4.6 -1138.4 (167.6) 1520.7 382.3 (77.6) (184.7) Cladonia- Mosses 13.5 -4.6 -634.1 828.1 194.1 (92.6) (31.5) (97.9) Brash 13.5 -7.5 0 819.3 819.3 (57.7) (57.7)

T5 = Soil temperature at 5 cm depth in °C, WT = water table depth in cm. NEE = net ecosystem exchange, PG = gross photosynthesis production, RTOT = ecosystem respiration

4.5 Discussion

4.5.1 Controls of CO2 fluxes

Peatlands respond to changing conditions of water table and soil temperature, switching from C sinks to sources (Alm et al., 1997; Alm et al., 1999a), or as evidenced in this study, to smaller or larger sources throughout the year. In our study, a positive relationship was determined between WT and soil temperature at 5 cm and CO2 although not as strong as that found in other studies (Maljanen et al., 2004; Riutta et al., 2007a; Renou-Wilson et al., 2014). Temperature dependence of

RTOT has been described in many studies (Silvolia et al., 1996; Bubier et al., 1998; Lafleur et al., 2005). Microbial activity is stimulated at higher temperatures promoting higher CO2 emissions (e.g. Frolking and Crill 1994; Silvola et al. 1996). Temperatures were high during the period of the study providing optimal conditions for soil respiration. The main fundamental requirements for C sequestration in previously degraded peatlands are the recolonization by vegetation of the rewetted

118 site and reinstatement of hydrological conditions similar to natural peatlands (Smolders et al., 2003). Despite maintaining a relatively high WT throughout the year, (mean depths > 16 cm), the study site remained a source of CO2-C, suggesting other governing factors are at play. Water table regulation of peat respiration is a contentious issue and other studies have questioned or rejected its validity (Lafleur et al., 2005; Dimitrov et al., 2010). However, it is accepted that WT position is one of the most critical factors affecting species composition of rewetted peatland vegetation communities (Maanavilja et al., 2014). Establishment of a stable and high WT is vital for recovery of vegetation at rewetted sites, although it may take years for ecosystems capable of peat accumulation to develop (Gorham and Rochefort, 2003). Water tables have remained high for the duration of this study and quite stable, although at individual brash collars fell to deeper depths. Recolonization of the rewetted site by peatland species is a reflection of WT dynamics on the site.

Previous authors have highlighted the influence vegetation has on CO2 sequestration in a peatland; vegetation cover, species composition and successional stage which a site exhibits determining whether it is a sink or source (Samartitani et al., 2011; Urbanova et al., 2012).

Vegetation on Scohaboy remains patchy and despite high levels of Sphagnum papillosum and Sphagnum capillifolium in the furrows, the site is in the early successional stages limiting its sink function capability. Winter CO2 loss rates are not yet being compensated for with summer PG, contributing to the large C source. Although the study took place soon after rewetting, a clear change in the vegetation was observed over the course of the study where Eriophorum vaginatum cover was spreading rapidly, as was that of the Sphagnum spp., particularly in the wetter areas of the site, similar to that observed by Komulainen et al., (1999). Eriophorum vaginatum has been shown to influence C sequestration due to both its high respiration rate and high rate of PG throughout the growing season (Tuittila et al., 2000b; Marinier et al., 2004). It quickly grows new leaves in early summer and photosynthesising biomass continues to grow through to late autumn with some leaves overwintering (Robertson and Woolhouse, 1984). Sedges have been identified as small C sinks (Minke et al., 2015), as well as vegetation mosaics of Eriophorum vaginatum and Sphagnum mosses in sites rewetted longer than Scohaboy (Kivimäki et al., 2008). Based on this evidence, the author is optimistic as regards the future

119 sink potential of the site. Initial signs of vegetation and WT dynamics support the view that the peatland has the potential to recover some C sink potential. Long term monitoring of the site is required to track changes which may take decades to occur (Bonnett et al., 2009), particularly in light of peatland sensitivity to weather conditions and climate change (Roulet et al., 2007; Bridgham et al., 2008; Urbanova et al., 2013).

4.5.2 Temporal variation in CO2 exchange

The seasonal variation in C flux components which we both observed and modelled (Figure 5.2, 5.3, 5.6) is well documented in Irish temperate conditions (Laine et al., 2007a; Wilson et al., 2007; Wilson et al., 2013). In the study period, fluxes were highest in June, July and August when the greatest biomass and highest PPFD levels are present on the study site (Figure 5.1). The seasonal changes in PG are mostly controlled by PPFD and GAI rather than temperature which control RTOT (Bubier et al., 1998; Griffis et al., 2000). Fluxes were smallest in the winter months at the coldest period of the year. Soil respiration continued and plant respiration was reduced as some plants senesced in the winter. Photosynthesis continued throughout the winter as was found by Sottocornola and Kiely (2005) due to the presence of evergreen species such as Sphagnum capillifolium and Eriophorum vaginatum on Scohaboy (Figure 5.6). The contribution of wintertime (November-March) photosynthesis was 11% of annual photosynthesis, smaller than respiration which totalled 17%. The winter RTOT flux observed here are within the range reported by Laine et al., (2007a) in a temperate natural peatland, and similar to those reported by Leppälä et al (2011a) in boreal conditions. High winter respiration levels are expected in the temperate climate due to the high temperatures experienced.

4.5.3 Carbon balance of Scohaboy Bog

The results of this study clearly show that rewetting does not produce immediate results. It is a long term strategy to reinstate the C sink function of a peatland and dependent on a high WT and peat forming vegetation on the site (Drösler et al., 2008). The results of this study show that previously forested peatland, three years -2 - following rewetting, remains a strong carbon source (585.3 ± 241.5 g CO2- C m yr

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1). Annual fluxes remain similar to those of drained, clear-felled sites as estimated by -2 -1 -2 -1 previous studies; 248-515 g CO2- C m yr and 207-539 g CO2- C m yr (Mäkiranta et al., 2007; Minkkinen et al., 2007) and much higher than a rewetted forestry site in Finland (Komulainen et al., 1999). Although Scohaboy has revegetated quickly, the ability of the vegetation to fix carbon cannot compensate for the high RTOT caused by the availability of highly decomposable fresh organic matter, similar to the situation in drained clear-felled peatland sites (Mäkiranta et al., 2012).

High CO2 emissions following rewetting have been found in other categories of rewetted peatland. Several studies have reported on the C source or sink strength of rewetted and restored peatlands with conflicting results owing to diverse restoration techniques, environmental conditions throughout the study year and time passed since beginning restoration (Tuittila et al., 1999; Bortoluzzi et al., 2006; Wilson et al., 2007b; Waddington et al., 2010; Strack et al., 2014). Carbon sink function returns within ten years on some peat extraction (Wilson et al., 2013; Strack et al., 2014) and agricultural (Knox et al., 2014) sites while others persist as C sources (Järveoja et al., 2016) even thirty years post rewetting (Petrone et al., 2003; Samartitani et al., 2011), despite re-establishing a high water tables, however older rewetted sites have been found to be carbon sinks (Samaritani et al., 2011), highlighting the unpredictability of re-establishing a C sink. Initial starting conditions on rewetted sites vary in vegetation, hydrological functioning, past management, pH and length of drainage (Höper et al., 2008; Rydin and Jeglum 2013), all of which influences their return to a C sink.

Carbon fluxes were surprisingly high in the sample plots in the plough furrows (Eriophorum- Sphagnum and Eriophorum) despite maintaining water tables no lower than 15 cm depth. Saiz et al., (2006) similarly found that within a forest stand, higher soil respiration rates were found in the furrows, attributable to a higher layer of organic matter gathered in the furrows than on the ridges or flat areas. The raised WT on our site is expected to reduce peat decomposition while fine roots and the easily decomposable humus organic matter layer decompose creating an additional release of C during the initial phase of restoration, expected to last 2-10 years (Höper et al., 2008). While sources across the whole year, during daylight hours when photosynthesising, these sample plots were sinks of CO2. Promisingly for restoration

121 both the Eriophorum and the Cladonia- Mosses vegetation communities acted as a small sink in April 2014. Annual fluxes were lower in the Cladonia- Mosses plots, when compared to the other sample plots, although NEE remained positive at all other times during the study period. As expected, the sample plots containing brash were the largest source of CO2 to the atmosphere, similar to Mäkiranta et al., (2012).

While small fluxes of PG were captured during onsite measurements, these were much smaller in magnitude than the vegetated plots and no relationship with any of the measured variables could be determined. It was assumed that values obtained during measurement were the result of unseen and undocumented mosses and so were not modelled. Therefore there may be a slight overestimation in the NEE values produced for the brash collars.

Aside from gaseous losses of C to the atmosphere, such as those in this study, C is also lost through waterborne carbon fluxes such as dissolved organic carbon (DOC), particulate organic carbon (POC), dissolved inorganic carbon (DIC) and gaseous

CO2 and CH4 , in run-off from the peatland (Evans et al., 2015). Surface waters from peatland catchments, saturated with CO2 and CH4 represent a significant pathway connecting C store with the atmosphere (Dinsmore et al., 2009a). Waterborne C can greatly influence the C balance of a peatland, turning it from a sink or neutral peatland to a carbon source (Roulet et al., 2007). Therefore, to accurately account for all C losses in future work, fluvial C losses should be measured.

4.5.4 Effect of brash on CO2 emissions

-2 Annual average C fluxes of the brash plots was almost double (819.9 g CO2- C m ) -2 that of the plots without brash (444.9 g CO2- C m ). This suggests that brash retention on a rewetted site significantly increases the CO2 emissions in the early years following rewetting. Brash decay generates CO2, however less than 40% of the increased fluxes are attributable to decomposition of brash on the site (Mäkiranta et al., 2012). Increased CO2 emissions could be attributed to the ‘priming effect’ i.e. the decomposition of peat beneath the brash could be increased by the introduction of fresh organic material from the brash. Studies have previously shown (Fontaine et al., 2004; Fontaine et al., 2007), the presence and significance of priming in organic matter decomposition. This affect appears to be most prominent where old soil organic matter, containing recalcitrant compounds and a low energy content, is

122 exposed to large amounts of fresh organic matter (Fontaine et al., 2004). Under natural conditions, energy from this old peat cannot maintain microbial activity. The introduction of fresh material provides microbes with an energy source, allowing them to further decompose the peat (Mäkiranta et al., 2012). In a rewetted forestry situation, the peat has not been removed; only disturbed and many layers of fresh organic material such as needles and small branches remain, providing large, fresh substrate for the microbes present in the peat causing a continuation of high gas emissions. Brash material remaining on the site has been found to affect nutrient content, particularly raising the potassium content (Palviainen et al., 2004; Hytönen and Moilanen, 2015). Increased nutrients will contribute to the priming effect, increasing microbial activity. Brash removal prior to rewetting is recommended to reduce the availability of fresh organic material, limiting the priming effect. High annual fluxes as reported here would be reduced and the C sink of the peatland would potentially be reinstated sooner should the brash material be removed.

4.6 Conclusions Our study highlights that peatland restoration can be a long term process in rewetted forestry determined by conditions present in the peatland prior to vegetation and the volume of fresh organic matter left on the site following clear-felling. The smallest

CO2 emissions were measured in the driest plots located on the ridges where a lower quantity of organic material would be collected supporting the evidence that fresh organic material creates a priming effect on the underlying peat, outweighing the effect of the high water table thereby increasing C emissions. Recolonization by peatland vegetation so soon after rewetting has created small localized sinks under particular conditions suggesting that in time, when the brash material and organic layer on the forest floor decomposes, the site will become a carbon store.

It is recommended that in rewetted forestry projects, clear-felled trees should be removed to reduce the volume of fresh organic material on the site encouraging a faster return of the C sink function. Further study should also be conducted on rewetted forestry on peatland sites sometime after rewetting to monitor the development of the site as it returns to a C sink.

Acknowledgements This project is funded by the Environmental Protection Agency (2012-CCRP- PhD. 2). Grateful thanks to Coillte for access to their LIFE 02 and LIFE 09 sites. Thanks

123 to Michael Kenny (Carlow Institute of Technology), Michael Kenna (Noone Engineering, Rathangan, Co Kildare) and Ray Byrne for assistance in equipment construction and David Wilson and Flo Renou Wilson for assistance with field measurements and analysis of results.

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Chapter Five

Methane and Nitrous Oxide flux dynamics in a blanket peatland forest eight years after rewetting

125

126

5.1 Abstract

Several forestry drained peatlands in the temperate region are currently undergoing rewetting attempts to return these ecosystems to their natural state. Drainage reduces methane (CH4) and increases nitrous oxide (N2O) emissions. Effective rewetting would not only return natural hydrological conditions and peatland vegetation but also trigger the resumption of CH4 emissions as microbial populations return and a reduction in N2O emissions. Gas emissions were measured on a rewetted blanket peatland eight years following rewetting using the chamber method. Micro sites comprising the dominant vegetation on the study site were established and measured -2 -1 for one year. The annual CH4- C balance was 2.94 ± 1.03 g CH4-C m yr and N2O balance was -11.78 μg m-2 year-1. Molinia caerulea plots were the largest sources of

CH4 followed by the Eriophrium vaginatum- Sphagnum plots. Great spatial variability was seen between the microsites. No relationship was found between N2O and any measured variable. More research is needed to track the progress of rewetted peatland forests as time since rewetting increases.

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128

5.2 Introduction

Natural peatlands are important C sinks and sources of both CH4 and low levels of

N2O (Minkkinen et al., 2002; Turunen et al., 2002). Natural wetlands, which include peatlands, account for 80% of natural CH4 emissions (Ciais et al., 2013). Although peatlands contain approximately 30% of the world’s soils organic nitrogen reserve

N2O fluxes are very low or insignificant in natural peatland (Martikainen et al., 1993; Regina et al., 1996; Minkkinen et al., 2002). Anaerobic methanogenic bacteria produce CH4 in the anoxic layer of the peatland as the last product of organic matter decomposition. They primarily reduce CO2 with molecular hydrogen and dismutate acetate (a root exudate), to CO2 and CH4 (Whalen 2005; Conrad, 2007; Nazaries et al., 2013). Aerobic methanotrophic bacteria present in the near surface oxidise CH4

(Whalen, 2005), consuming CH4 as a carbon and energy source (Murrell 2010). Methane is released from peat by ebullition, diffusion or plant facilitated transport (Joabsson et al., 1999; Lai, 2009). Methane emissions are influenced primarily by soil temperature and water table (Dunfield et al., 1993; Komulainen et al., 1998; Joabbson and Christensen, 2001). Methane fluxes on peatlands are also connected with peat aeration, nutrient level, vegetation cover (particularly the presence of aerenchymatic species), peat compaction and pH (Williams and Crawford, 1984; Diese, 1993; Nykänen et al., 1998; Ström et al., 2005; Armstrong et al., 2015).

Plants have a direct effect on CH4, by altering peat bacteria community structure (Robroek et al., 2015) and particularly plants such as Juncus effusus and Eriophrium vaginatum (Henneberg et al., 2015; Mariner et al., 2004) influence emissions by providing a direct route for the release of CH4 to the atmosphere, through aerenchyma, thereby avoiding the methanotrophic bacteria (Whalen 2005; Green and Baird, 2011).

Nitrous oxide is produced in both nitrification (aerobic) and denitrification (anaerobic) processes by bacteria in the soil (Davidson and Schimel, 1995).

Denitrification represents the most important N2O source in peatlands (Regina et al., 1995) and principally involves the bacterial genera Pseudomonas, Micrococcus, Bacillus, and Thiobacillus in which carbon functions as the source of energy and nitrogen the electron acceptor (Kamal and Varma, 2008). During the denitrification process, nitrogen is reduced to the gases N2O and N2 (Zumft, 1993; Ye et al., 1994).

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+ Nitrous oxide may also be produced during nitrification through oxidation of NH4 by aerobic chemolithotrophic bacteria (Koops et al., 1997; Martikainen et al., 2002).

Drainage of peatlands reduces CH4 (Sundh et al., 2000) while increasing N2O emissions (Minkkinen et al., 2002). Following rewetting, CH4 emissions increase due to the raised water table, N2O levels are low and vegetation cover increases (Tuittila et al., 2000; Beyer and Höper, 2014). Despite this, when balanced against their C sink potential, peatlands are potentially valuable climate change mitigation tools. Rewetting is perceived as a practical means of restoring the carbon sink function of peatlands (Höper et al., 2008). Rewetting involves drain blocking (Bragg 2011) typically followed by other interventions (Quinty and Rochefort, 2003; Wheeler, 1995) to limit further peat degradation (Schuman and Joosten, 2008) and re-establish an ecosystem similar to the one which was drained (Konvalinková and Prach, 2014). In the future, due to climate change, rewetting and management of peatlands will become more complex (Erwin, 2009; Grand Clement et al., 2013). Thus extensive knowledge of peat GHG dynamics is essential in the adaptation of management regimes and consequentially will improve the effectiveness of any rewetting. However data about the C balance of rewetted peatlands is scarce in the temperate climate zone (e.g. Hendricks et al., 2007; Renou-Wilson et al., 2016; Wilson et al., 2013; Wilson et al., 2016b) and particularly studies on rewetted peatland forestry are lacking (e.g. Komulainen et al., 1999), therefore research is urgently needed.

We studied the CH4 and N2O balances from a rewetted blanket peatland forest in the temperate region. Our main goal was to investigate the effect restoration had on the dynamics of CH4 and N2O, determine had CH4 emissions returned to similar levels of natural peatlands and to establish the governing variables of CH4 and N2O.

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5.3 Materials and Methods

5.3.1 Site Description

See Section 3.3.1 for site descripton

5.3.2 Study Site

See section 3.3.2 for study site descripton

5.3.3 Measuring CH4 and N2O fluxes

Methane sampling was conducted at fortnightly intervals from March 2014 to June 2015 except for one month period December 2014- January 2015. Fluxes were measured using the closed chamber method, adapted from that used by Regina et al. (1996). A 60 cm × 60 cm × 30 cm opaque polycarbonate chamber was placed on the collar. Each chamber had a vent ensuring pressure equilibrium which was only closed after the chamber had been placed on the collar. The chamber was fitted with a fan which circulated air inside the chamber. A water filled channel at the top of the collar created an air tight seal during sampling. Four 50 ml samples were taken, normally at 5 minute intervals. A 10 minute interval was used in winter months when low flux rates occur due to the colder temperatures and low plant cover. Gas samples were taken in plastic syringes fitted with stopcocks and transferred to pre- evacuated glass vials (Code 839W, Labco Ltd, UK) for transport to the laboratory for analysis on a gas chromatograph.

Samples were analysed for CH4 and N2O at Justus Liebig University Giessen, Germany within 2 months using a gas chromatograph (Bruker Greenhouse Gas Analyser 450-GC) fitted with a thermal conductivity conductor (TCD), a flame ionisation detector (FID) and an Electron Capture Detector (ECD). Detector temperatures were 200°C (TCD), 300°C (FID) and 300°C (ECD) and the oven -1 temperature was 70°C. Nitrogen was used as the carrier gas (22 ml min ). The CH4

(1.02, 1.81, 5.02, 20.9 and 100.1 ppm) and N2O (0.248, 0.321, 2.01, 15.1 and 100.1 ppm) standards were supplied by Deuste Steininger GmbH. Gas concentrations were

131 calculated using the Galaxie software (Varian Inc., 2006). Fluxes (mg m-2 h-1) were calculated from the linear change in gas concentration as a function of time, chamber volume, collar area and air temperature. A flux was accepted if the coefficient of 2 determination (r ) was at least 0.90. Positive values indicated losses of CH4 and N2O to the atmosphere, and negative flux values indicated CH4 and N2O uptake.

Figure 5.1 CH4 and N2O equipment set up. Clockwise from top left: 1- 60 cm x 60 cm collar inserted into the peatland with water table measuring pipe located beside it, 2- chamber equipped with battery operated fan inside. Thermometer is used to measure air temperature inside the chamber, 3- chamber stoppered during sample collection, samples are injected into glass vials, 4- Samples are collected with a plastic syringe.

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5.3.4 Environmental Variables

Air temperatures inside the chamber and in the peat at 5 cm and 10 cm (T5, T10) depths were measured simultaneously at sampling time. Soil temperatures were measured using a temperature probe (Jenway 220). Air temperature was measured using a Ted Pella Inc., 28163 Traceable Total-Range Thermometer. The level of the water table was measured using PVC pipes with a series of small holes pierced at regular intervals along the side. These pipes were positioned beside the collars prior to the start of the study, leaving 10cm of piping visible above the ground. Water table depth was measured using a Pocket Dipmeter Kill Mini 10 m in length supplied by HYDROKIT Dorset. The water table data was interpolated between measurements in order to describe water table patterns across the seasons. GAI was estimated by summing the green area (GA) of each species as described in section 3.3.5. A meteorological station (Watchdog 1400 Micro station) was installed on each site measuring PAR and soil temperature at 5 cm and 10 cm at ten minute intervals throughout the duration of the study.

5.3.5 Flux estimation and statistical analysis

As stated previously, four samples were taken at 5 or 10 minute intervals at each collar. In cases where there was ebullition or sampling error during the measurement period, indicated by a high CH4 concentration unrelated to the others in that measurement period, the measurement was discarded. Similarly, if ebullition was apparent at the second or third sample, the measurement was rejected. If ebullition was apparent only at the end of the sample period (fourth sample), the last sample was discarded and the flux calculated using the first three samples. Therefore, the data utilised include only the diffusive or plant mediated transport emissions of CH4 from the peat. Methane emitted by ebullition was not included as it is possible that the ebullition was caused by the sampling method itself. A maximum of one discarded sample was permitted per measurement. If vial leakage was evident the sample was also discarded. From a total of 1024 samples, 97% were deemed acceptable.

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Plots were grouped together following the relationships determined when modelling

CO2 flux components in Chapter Four and based on similar vegetation patterns (Fig 5.1). The first group is comprised of sample plot P1 containing Sphagnum sp and Eriophorum sp, the second group is P5 which although Molinia was present in the plot, it was the only plot to contain Calluna vulgaris and Cladonia portentosa and behaved uniquely to the remaining plots when modelling. Group three is made up of the remaining collars, P2, P3, P4, P6, P7 and P8. Molinia caerulea is dominant in these collars with other moss species occurring in the undergrowth.

Table 5.1 Vegetation species recorded in the study plots. Species are listed in descending order of dominance.

Flux estimates were calculated for each vegetation group over the sampling period. Annual estimates were determined by linear interpolation between the measured data throughout the study period as per Urbanová et al. (2013). No relationship was found between N2O emissions and either soil temperature or water table depth. Seasonal means of N2O were calculated to determine approximate annual emissions.

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5.3.6 CH4 and N2O flux modelling

No relationship between any of the measured variables and either gas was established during this study. Therefore regression modelling was not used.

Estimates of annual balances for CH4 were instead determined using linear interpolation. Nitrous oxide daily emissions were estimated using seasonal means. Interpolation was deemed inadequate due to the number of missing values from sample days.

5.4 Results

5.4.1 Environmental variables

In both study years, annual rainfall was greater than the 30 year average of 1434.97 mm (Figure 5.2 a). In year 1, 1657.5 mm (16% wetter) and in year 2, 1764 mm (23% wetter) rain was recorded. In both years highest rainfall was recorded in the November-February period and the lowest in the summer months, as is typical for a temperate climate. In 2014, maximum air temperature at Athenry Meteorological Station was 26.6 °C recorded in July 2014 and minimum air temperature was -4.7 °C in December. The following year showed a similar pattern, maximum air temperature recorded at the end of June (22.8 °C) and a minimum of -5.4 °C recorded in February. Maximum soil temperatures in both years were reached in July 2014 and 2015 and minimum during January and February (Fig 5.2b). Water table depth remained between 1 cm and 56 cm below the peat surface (Fig 5.3). All sample plots experienced a fall in water table during the summer months. The largest decrease was observed in P7 a Molinia caerulea dominated plot, in October 2014, where WT fell to -56 cm. The lowest, mean annual WT was observed in P7 in both 2014 (-27 cm) and 2015 (-19 cm), while the highest was observed in P3, in 2014 (-7 cm) and 2015 (-6 cm), also a Molinia caerulea dominated sample plot. The C/N ratio from the soil profile, 0- 80 cm was 34.4.

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400 (a) Year 1 300 Year 2 200

100 Precipitation (mm) Precipitation

0 Jan Feb March April May June July Aug Sept Oct Nov Dec

25 (b) 20

5 Soil temperature (°C) temperature Soil

0 Mar Jul Nov Mar Jul Nov Mar 2014 2015 2016

Figure 5.2 Climate data for Pollagoona, Co Clare, during the sample period. (a) monthly average rainfall 2014- 2015 (mm) (Met Eireann, Gort rainfall Station), and (b) soil temperature (°C) at 5 cm depth from March 2014- March 2016. Dark circles and line indicate 30 year average (1984-2014, www.met.ie).

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-10

-20

-30

-40

-50

Water table depth (cm) depth table Water -60 Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan Feb Mar Apr May Jun Jul 2014 Month 2015 Figure 5.3 Interpolated water table depths in all study plots from 3rd March 2014 to June 30th 2015.

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5.4.2 Measured gas fluxes

(a)

(b)

(c)

2014 2015

-2 -1 Figure 5.4 Observed CH4 fluxes (mg CH4 m h ) within the (a) Eriophorum vaginatum – Sphagnum, (b) Cladonia portentosa- Calluna vulgaris and (c) Molinia dominated communities at Pollagoona Co. Clare. Positive values indicate a loss of

CH4 to the atmosphere and negative values indicate CH4 uptake by the peat. Note differences in scale on the y- axis.

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Measured CH4 fluxes followed the sequence of Molinia > Cladonia portentosa- Calluna vulgaris > Eriophorum vaginatum- Sphagnum plots (Fig 5.4). Molinia caerulea plots had a mean water table depth of 12.5 cm. The least variation in fluxes and lowest mean water table (-18.9 cm) were observed in Cladonia portentosa- Calluna vulgaris vegetation communities. Fluxes are lowest in plots P5, P6 and P7 with an average water table depth of 21.6 cm. Although no significant relationship between water level and CH4 emissions could be determined, water table depth appears to be influencing flux levels as emissions follow WT changes. No flux occurred intermittently in three sample plots, between March and May 2014 and in one plot between November 2014 and February 2015. Methane fluxes varied temporally across the seasons. Methane emissions peaked in the summer months of June July and August before falling steadily in the autumn months. Average sample -2 -1 day measured methane emissions varied from -0.00125 mg CH4 m hr to 1.29 mg -2 -1 CH4 m hr . The mean daily sampled growing season emissions were 46.5% greater than winter emissions. All sample plots except one acted as a CH4 sink in May 2014. The highest emissions were recorded on 28 July 2014 in a Molinia dominated plot -2 -1 (3.93 mg CH4 m hr ) while the same plot acted as the greatest sink on 15 April -2 -1 2014 (0.24 mg CH4 m hr ). Emissions in June and July of 2014 were greater than those in 2015, most likely due to the slightly lower soil temperatures and higher WT in 2015. The significance threshold used in this study was p≤ 0.05. One way ANOVA test confirmed a significant difference between sample plot means (p<

0.05). A significant but weak relationship was found between the measured CH4 emissions and soil temperature (r2= 0.20, p< 0.05), while no correlation was found between CH4 fluxes and either water table or GAI.

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(a)

(b)

(c)

2014 2015

-2 -1 Figure 5.5 Measured N2O fluxes (μg m hr ) within the (a) Eriophorum vaginatum – Sphagnum, (b) Cladonia portentosa- Calluna vulgaris and (c) Molinia dominated communities at Pollagoona Co. Clare. Positive values indicate a loss of N2O to the atmosphere and negative values indicate N2O uptake by the peat. Note differences in scale on the y-axis.

Measured N2O fluxes over the study period were variable with both positive, negative and zero values recorded, although all were small (S.D = 0.05). Peaks in

N2O emissions were apparent in the summer period (Fig 5.5). For the Eriophorum vaginatum- Sphagnum plot, these may be connected to periods of low water table. No such association is apparent for Cladonia portentosa- Calluna vulgaris plots. The range of sample day emissions in Molinia plots was much greater in the summer

140 months 2014 and into November, than in winter 2014- 2015. As temperatures rose in spring and summer 2015, the range of sample day emissions once again began to increase. Average sample day emissions ranged between 0.0014 and 0.054 μg m-2hr- 1.

5.4.3 Calculated Gas fluxes

1.0

0.8

0.6 - C month C - 4

g CHg 0.4

0.2

0.0 March April May June July Aug Sept Oct Nov Dec Jan Feb 2014 Month 2015

-2 -1 Figure 5.6 Interpolated monthly means of CH4-C (g m month ) within the Eriophorum vaginatum- Sphagnum, Cladonia portentosa- Calluna vulgaris and Molinia caerulea dominated communities at Pollagoona Co. Clare. Error bars represent standard deviation on the interpolated means.

Despite finding no correlation between CH4 fluxes and either WT or GAI and only a weak relationship with soil temperature, fluxes appeared to follow the seasonal trends of all three variables (Fig 5.6). Fluxes tended to peak in mid-summer, when soil temperature and GAI was highest and WT lowest. Spatial variation occurred across the sample plots, caused by varying WT, vegetation composition and the presence or absence of brash. Water table levels varied between 1 cm and 28 cm

141 below the peat surface in sample plots over the study period. Average sample period water levels ranged from -5.5 cm to -28 cm depth. Water table levels remained below -10 cm in all sample plots in June and July 2014, sometimes dropping below -

20 cm, and below -10 cm in four plots from April- August 2014. Low CH4 emissions were recorded in May 2014 causing the interpolated balance to remain below 0.1 g CH4- C in the month. Methane emissions in the course of this study followed the water gradient, being highest where the mean annual WT was above 10 cm while lowest emissions were recorded in plots whose mean annual WT was approximately -20 cm, in line with suggestions by Couwenberg et al., (2011).

Table 5.2 Summary of annual sums of CH4 and N2O and averages of relevant variables from the three study microsites. Standard errors of the estimates are shown in brackets for all fluxes. Positive values indicate a loss of C or N from the site and negative values indicate an uptake of C or N to the site. T5 = soil temperature at 5cm depth in °C, WT = water table depth in cm.

T5 WT CH4-C N2O (g m-2 yr) (μg m-2 yr-1) Eriophorum/ 12.6 -10.2 1.99 88.46 Sphagnum Cladonia/ Calluna 12.6 -18.9 0.54 -169.50 Molinia 12.6 -13.4 2.53 (1.01) 45.70 (0.09)

-2 -1 An annual balance of 2.94 ± 1.03 g CH4-C m yr was calculated for Pollagoona by interpolation of measured emissions and averaging annual balances of microsites (Table 5.2). High levels of variance are evident both temporally within and spatially between the sample plots. All plots were CH4 sources for the study period, with -2 -1 interpolated annual emissions ranging from 0.081g CH4-C m yr (P7) to 8.24g -2 -1 CH4-C m yr (P4). The N2O annual balance as estimated by seasonal means was - 11.78 μg m-2 year-1. Great variation was seen between the three microsites with the

Cladonia portentosa- Calluna vulgaris site acting as a N2O sink for the duration of the study while the remaining microsites were sources throughout the same period.

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5.5 Discussion

To date, no other study has been conducted which quantifies the CH4 fluxes from a rewetted blanket peatland forest. The results demonstrate that such sites remain low -2 -1 emitters of CH4. Estimated annual emissions of 2.21± 1.03 g CH4-C m yr are -2 -1 much lower than 14.46 g CH4-C m yr calculated from values reported by Koskinen et al., (2016) from rewetted spruce swamps in Finland but within the range -2 -1 (0.082 – 2.96 g CH4-C m yr ) estimated from Juottonen et al., (2012), also on rewetted forests in Finland, although the reported mean WT was much higher than in this study. Much greater CH4 volumes have been reported in previous studies from rewetted industrial extraction sites both in Ireland (Wilson et al., 2009) and in

Canada (Mariner et al., 2004). The CH4 flux values recorded here are lower than 6.2 −2 −1 g CH4 m year reported for an Atlantic blanket bog in Kerry, south west Ireland (Laine et al., 2007b).

At all times during this study, the WT remained below ground level, maintaining aerobic conditions in the upper peat layers. Given that the area of highest methanogenic activity is normally located close to the water table, at the boundary between the aerobic and anaerobic layers (Segers, 1998); deep water tables increase the distance to travel from gas production to release. In addition, the presence of oxygen in the aerobic zone above the WT restrains the action of the CH4 producers

(Nykänen et al., 1998), limiting CH4 emissions. If the WT on Pollagoona increases in the future, CH4 emissions would be expected to rise accordingly. Although our emissions are larger than drained sites, which are reported to emit as little as < 1g -2 -1 CH4 m yr (Nykänen et al., 1996; Penman et al., 2003; Maljanen et al., 2004;

Renou- Wilson et al., 2014); this study indicates that while CH4 emissions are expected to increase following rewetting (Komulainen et al., 1998), on rewetted forest sites the CH4 emitted continues to be less than from rewetted industrial sites and similar or lower than natural sites eight years after rewetting, increasing their potential as climate change mitigation tools when leveraged against possible C storage capacity.

Little brash was present in the sample plots as eight years following rewetting, it had decomposed. Plots containing brash were included in those with a WT > -10 cm so

143 although Mäkiranta et al., (2012) reported that brash plots emitted more CH4 than control plots, it cannot be definitely said that the higher emissions in this study can be contributed to the presence of brash.

Drainage is known to influence peatland microbial diversity and composition (Fenner et al., 2005: Jaatinen et al 2007). In addition to relatively low annual mean

WT, remaining historical influences of drainage could be contributing to low CH4 emissions. Following rewetting, methanogen numbers increase and methanogenic community configuration becomes more similar to natural peatlands (Urbanová et al., 2011), however, the numbers of both bacteria groups remain lower than in natural sites (Juottonen et al., 2012) limiting CH4 production in rewetted peatlands as is evidenced in this study by the lower than natural peatland volumes of CH4 emissions.

Successful restoration initiates vegetation succession towards that of natural peatland (Tuittila et al., 2000a; Komulainen et al., 1999). It has been established that vegetation control of CH4 emissions is subsequent to WT control (Leppälä et al., 2011); therefore it will only influence emissions when WT is at a suitable depth to the soil surface. Despite this, vegetation composition of a rewetted peatland has been found to influence the C balance, both for its CH4 transport capability and decomposing litter quality (Samaritani et al., 2011; Vanselow-Algan et al., 2015;

Zak et al., 2015). Increased CH4 emissions in this study, compared to drained sites, is assisted by the colonization of Eriophorum and Juncus, which facilitate CH4 transport, (Carol and Freeman, 1999; Mariner et al., 2004; Henneberg et al., 2015b), and also potentially influence microbes involved in CH4 dynamics (Robroek et al.,

2015). Eriophorum and Juncus most likely contribute to the higher CH4 emissions observed in sample plots P1 and P4, when compared to other plots. Despite aerenchymous species being present, CH4 emissions are not abnormally high. While commonly associated with CH4 transport and release, it cannot be forgotten that the primary function of aerenchymous species is to transport oxygen to their waterlogged roots. It has been observed in previous studies that the oxygen supply is large enough to not only supply the roots but also oxidise the surrounding soil and produce an oxidised rhizosphere (Mitsch and Gosselink, 2007), likely being detrimental to CH4 production. Sphagnum, despite not being aerenchymous, has been documented as facilitating CH4 transport (Roura-Carol and Freeman, 1999),

144 thereby promoting emissions from sample plots P3, P5 and P7. Molinia caerulea has been linked to low emission rates in some instances (Bhullar et al., 2014), while high in others (Vanselow-Algan et al., 2015). It has been suggested that Molinia contributes to CH4 emissions due to the large amounts of litter produced by Molinia caerulea which is more easily decomposed than other species found on site (van Breemen, 2013; Vanselow-Algan et al., 2015). Despite these plots having the highest emissions in this study, they remain low when compared to studies mentioned previously and it can be suggested, given the low CH4 fluxes in the Molinia caerulea dominated plots that Molinia is most likely not acting to boost emissions. During the growing season, when Molinia caerulea was abundant, gas emissions increased compared to winter emissions, however this can be contributed to additional factors, such as soil temperature and increased microbe activity, as well as the occurrence of this plant.

Similar to our study, other studies have suggested that there is little or no effect on

CH4 emissions one year after rewetting (Urbanová et al., 2012), however there is evidence that peatlands will emit growing volumes of CH4 as time passes since restoration; increases in CH4 fluxes have been observed within three years (Waddington and Day, 2007), as the peatland ecosystem returns to natural conditions. Methane emissions from restored peatlands may still remain less than those of natural sites (Strack et al., 2013), except in ditches on rewetted sites, which become CH4 hotspots (Waddington and Day, 2007). Pollagoona maintains a relatively low water table throughout much of the year and that, in addition to few flooded drains onsite, maintains conditions for low CH4 emissions. As succession continues and microbe populations recover, provided WT increases across the site, emissions may increase. Further investigation of rewetted peatland development in a temperate blanket bog is required to track the evolution of the peatland and subsequent emission levels.

Although interpolation of measured values is a common method of estimating CH4 emissions where regression modelling is not possible, it is likely to lead to underestimation in annual estimates. Based on measured fluxes on sampling days, ebullition events and short term periods of high emissions following rainfall may have been missed and therefore not accounted for.

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Nitrous oxide emissions were low but variable throughout the study period with no obvious temporal or spatial trend. Emissions in some sample plots appear to be related to WT drawdown, however, given the complexity of interactions governing

N2O and the limited variables measured in this study, it cannot be confirmed. Given the size of emissions recorded, it is possible that small minus and positive values are errors around zero. In addition, due to the sampling method used in this study (stationary sampling points visited every two weeks), we may have avoided ‘hot moments’ of emissions or ‘hot spots’ (Groffman et al., 2009; Butterbach et al., 2013) in the study site. The C/N ratio for this site is greater than 25, the threshold above which N2O emissions are negligible (Klemedtsson et al., 2005). Therefore, our small emissions are to be expected. Nitrous oxide emissions are governed by environmental factors such as temperature, substrate, redox conditions, pH, water table and competition from vegetation (Machefert et al., 2002; Müller and Sherlock, 2004; Silvan et al., 2005; Lohila et al., 2010; Butterbach et al., 2013). Emissions are expected to decrease rapidly after rewetting, falling to almost zero if water tables < - 15 cm (Couwenberg et al., 2011; IPCC 2014b).

Studies on temperate peatlands where N2O emissions have been measured are low

(Beyer and Höper, 2015; Vanselow-Algan et al., 2015) and N2O emissions have been assessed as being small or insignificant on natural peatlands (Martikainen et al., 1993; Regina et al., 1996; Alm et al., 2007). Following drainage for forestry these may increase (Minkkinen et al., 2002). Peatlands drained for forestry have been recorded as producing 0.03- 0.92 g m-2 yr-1 (Von Arnold et al., 2005; Ojanen et al.,

2010). Comparisons in Estonia found that N2O emissions from rewetted sites were a small fraction (- 0.12 – 2.13 μg N m-2h-1) of emissions from peat mining sites (27.1 -2 -1 μg N m h ) (Järveoja et al., 2016) highlighting the effect of rewetting on N2O fluxes. Studies conducted on rewetted blanket bog peat extraction sites in Ireland have not detected any N2O emissions (Wilson et al., 2013; Wilson et al., 2016b), and the IPCC considers N2O emissions negligible for rewetted sites tier 1 accounting (IPCC, 2014b). It can be argued that the low levels measured in this study represent the peatland returning to almost zero levels of emissions following drainage. Our results indicate N2O emissions may progress towards natural conditions quite quickly. The results of this study indicate that rewetted peatland forestry remains a small source of N2O eight years post rewetting.

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5.6 Conclusions

The findings of this study are important for large area of rewetted forestry in Ireland.

Eight years post rewetting, it can be seen that CH4 and N2O fluxes are similar to those from their natural counterparts. Methane emissions have most likely increased since rewetting while N2O emissions have become almost negligible. The restoration of a carbon sink function and pre-drainage hydrology following rewetting is a long term process, and in time, should the water table remain high, this site is expected to regain its carbon sink function. Should this happen, it provides optimism for the potential of the site to once again return to a natural functioning peatland. Increased monitoring of rewetted forestry sites in Ireland is required to track their progress and develop a greater understanding of vegetation development and changing gas dynamics.

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148

Chapter Six

Methane and Nitrous Oxide emissions in a raised bog after clearfelling and rewetting

149

150

6.1 Abstract

Methane (CH4) and nitrous oxide (N2O) are the most important GHGs following

CO2 and contribute greatly to the threat of global warming. Drainage disrupts the

GHG dynamics in a peatland, reducing CH4 and increasing N2O emissions. Several forestry drained peatlands in Ireland have undergone rewetting attempts to return the hydrological function of these ecosystems which in turn is expected to re-establish the C sink function and peatland vegetation. Gas emissions were measured on a raised peatland three years following rewetting using the chamber method. Micro sites comprising the dominant vegetation on the study site were established and -2 - measured for one year. The annual CH4-C balance was 3.25± 0.0.58 g CH4-C m yr 1 -2 -1 and N2O balance was 72 μg m year . Brash plots were the largest sources of CH4 followed by the Cladonia portentosa- Mosses plots. Great spatial variability was seen between the microsites. No relationship was found between N2O and any measured variable. Three years following rewetting CH4 emissions from raised peatland are similar to those from natural sites and N2O emissions are almost negligible. Further monitoring is essential to track emissions as succession continues.

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152

6.2 Introduction

Nitrous oxide and CH4 are the most important GHGs after CO2 which are exchanged between the atmosphere and terrestrial ecosystems. The global warming potential

(GWP) of CH4 is 25 and N2O 298 times higher than CO2 over 100 years (Forster et al., 2007). Generally, natural peatlands are regarded as being large sources of CH4

(Crill et al., 1992; Laine et al., 2007b), while N2O emissions are low (Regina et al., 1996; Juszczak et al., 2013). Wetlands, to which peatlands belong, globally emit between 80 and 280 Tg (Bartlett and Harris, 1993; Mathews, 1996; Bridgham et al.,

2013), supplying 20-30% of the global CH4 emissions (Houghton et al., 2001). Despite covering approximately 3% of the world’s land surface, natural peatlands store as much carbon as all terrestrial biomass (Parish et al., 2008). Methane fluxes are controlled by various factors; vegetation composition, water table depth, temperature, microbe composition, peat aeration, nutrient level, peat compaction, pH (Williams and Crawford, 1984; Diese, 1993; Nykänen et al., 1998; Ström et al., 2003; Ström et al., 2005; Juottonen et al., 2012; Armstrong et al., 2015; Henneberger et al., 2015a). Drainage for forestry significantly changes the C cycle in a peatland. Aerobic decomposition is increased as the water table drops (Blodau and Moore, 2003). Expansion of the aerobic zone above the water table inhibits methanogenic activity and at least at first encourages methanotrophs (Kettunen et al., 1999). Nitrous oxide is produced by nitrification (aerobic) and denitrification (anaerobic) processes (Davidson and Schimel, 1995). In highly anaerobic conditions, N2O can be consumed by denitrification. Once drained for forestry, peatland CH4 emissions decrease (Ojanen et al., 2010), with the exception of drainage ditches which may be hotspots of CH4 emissions (Sundh et al., 2000; Schrier-Uijl et al., 2008; Hyvönen et al., 2013) and N2O emissions increase (Von Arnold et al., 2005). Rewetting is accepted as a viable means of restoring the C sink function of peatlands (Höper et al., 2008), described as the intentional act of increasing the water table on drained soils to recreate water saturated conditions (IPCC, 2014b). Successful restoration from an ecological point of view is considered to be the re-establishment of peat forming vegetation (e.g. Rochefort et al., 2003), return of the C sink function to the peatland (Hendricks et al., 2007; Beyer and Höper et al., 2015) and actively accumulating organic matter (Lucchese et al., 2010), facilitated by hydrological conditions similar to those which existed prior to drainage. Methane emissions rise

153 following rewetting caused by increasing CH4 production and decreasing oxidation (Waddington and Day, 2007; Wilson et al., 2009), although emission levels may remain lower than natural sites (Tuittila et al., 2000b). Nitrous oxide emissions are reduced considerably on rewetted sites (Tauchnitz et al., 2015; Wilson et al., 2013; Renou-Wilson et al., 2016). Rewetting and especially monitoring of rewetted sites, is a relatively recent activity, the number of studies available in the temperate region are limited (e.g. Hendricks et al., 2007; Renou-Wilson et al., 2016; Wilson et al., 2013; Wilson et al., 2016b), particularly in peatland forestry (e.g. Komulainen et al., 1998; Komulainen et al., 1999; Juottonen et al., 2012). Extensive knowledge of peat GHG dynamics is essential in the adaptation of rewetting management regimes and to improve the effectiveness of any rewetting. In this project, we assessed the CH4 and N2O balances from a rewetted raised peatland forest in the temperate region and present our results as annual balances. Our objectives were to (a) measure CH4 and

N2O dynamics in rewetted raised peatland (b) determine whether CH4 and N2O emissions had returned to similar levels of natural peatlands (c) investigate the relationship of CH4 and N2O to environmental variables. We assess the need for further research in light of the probable continuation of rewetting in the future.

6.3 Materials and Methods

6.3.1 Site Description

See Section 4.3.1 for site description

6.3.2 Study Site

See Section 4.3.2 for study site description

6.3.3 Measuring CH4 and N2O fluxes

Methane sampling was conducted at fortnightly intervals from March 2014 to June 2015 except for one month period in December 2014- January 2015. Fluxes were measured using the closed chamber method, adapted from that used by Regina et al. (1996). A 60 cm × 60 cm × 30 cm opaque polycarbonate chamber was placed on the

154 collar. Each chamber had a vent ensuring pressure equilibrium which was only closed after the chamber had been placed on the collar. The chamber was fitted with a fan which circulated air inside the chamber. A water filled channel at the top of the collar created an air tight seal during sampling. Four 50 ml samples were taken, normally at 5 minute intervals. A 10 minute interval was used in winter months when low flux rates occur due to the colder temperatures and low plant cover. Gas samples were taken in plastic syringes fitted with stopcocks and transferred to pre- evacuated glass vials (Code 839W, Labco Ltd, UK) for transport to the laboratory for analysis on a gas chromatograph.

(1.02, 1.81, 5.02, 20.9 and 100.1 ppm) and N2O (0.248, 0.321, 2.01, 15.1 and 100.1 ppm) standards were supplied by Deuste Steininger GmbH. Gas concentrations were calculated using the Galaxie software (Varian Inc., 2006). Fluxes (mg m-2 h-1) were calculated from the linear change in gas concentration as a function of time, chamber volume, collar area and air temperature. A flux was accepted if the coefficient of 2 determination (r ) was at least 0.90. Positive values indicated losses of CH4 and N2O to the atmosphere, and negative flux values indicated CH4 and N2O uptake.

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Figure 6.1 CH4 and N2O equipment set up. Clockwise from top left: 1- 60 cm x 60 cm collar inserted into the peatland with water table measuring pipe located beside it, 2- chamber equipped with battery operated fan inside. Thermometer is used to measure air temperature inside the chamber, 3- chamber stoppered during sample collection, samples are injected into glass vials, 4- Samples are collected with a plastic syringe.

6.3.4 Environmental Variables

156 to the start of the study, leaving 10cm of piping visible above the ground. Water table depth was measured using a Pocket Dipmeter Kill Mini 10 m in length supplied by HYDROKIT Dorset. The water table data was interpolated between measurements in order to describe water table patterns across the seasons. GAI was estimated by summing the green area (GA) of each species as described in section 3.3.5. A meteorological station (Watchdog 1400 Micro station) was installed on each site measuring PAR and soil temperature at 5 cm and 10 cm at ten minute intervals throughout the duration of the study.

6.3.5 Flux estimation and statistical analysis

As stated previously, four samples were taken at 5 of 10 minute intervals at each collar. In cases where there was ebullition or sampling error during the measurement period, indicated by a high CH4 concentration unrelated to the others in that measurement period, the measurement was discarded. Similarly, if ebullition was apparent at the second or third sample, the measurement was rejected. If ebullition was apparent only at the end of the sample period (fourth sample), the last sample was discarded and the flux calculated using the first three samples. Therefore, the data utilised include only the diffusive or plant mediated transport emissions of CH4 from the peat. CH4 emitted by ebullition was not included as it is possible that the ebullition was caused by the sampling method itself. A maximum of one discarded sample was permitted per measurement. If vial leakage was evident the sample was also discarded. From a total of 974 samples, 97.5% were deemed acceptable.

The significance threshold used in this study was p≤ 0.05. One way ANOVA test confirmed a significant difference between vegetation categories (p= 0.001). Plots were grouped together based on dominant vegetation and the relationships determined in Chapter Four (Table 6.1). The first group is comprised of sample plot S1, S2 and S5, dominated by Sphagnum spp. and Eriophorum, the second group is S3 and S6 whose vegetation is primarily Cladonia portentosa, Calluna vulgaris and mosses, while group three is made up of the three brash collars, S4, S7 and S8.

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Table 6.1 Vegetation species recorded in the study plots. Species are listed in descending order of dominance.

Flux estimates were calculated for each vegetation group. Annual estimates were determined by linear interpolation between the measured data throughout the study period as per Urbanová et al. (2013). No relationship was found between N2O emissions and either soil temperature or water table depth. Seasonal means of N2O were calculated to determine approximate annual emissions.

6.3.6 CH4 and N2O flux modelling

No relationship between any of the measured variables and either gas was established during this study. Therefore regression modelling was not used.

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6.4 Results

6.4.1 Environmental variables

Annual rainfall in both study years was similar to the 30 year average of 948.2 mm (Figure 6.2 a). Precipitation in year 1 was 912.5 mm, 4% less and in year 2 was 1126.5 mm, 19% wetter than the average. The highest rainfall in both years occurred in the winter months and the lowest in June-July, as is typical for this climate. In year 1, mean air temperature at Gurteen Meteorological Station, was 9.9°C and in year 2, 9.5°C. Maximum temperatures in both years were recorded in July and minimum temperatures in February. Soil temperatures in both years reached their maximum in July 2014 and 2015 and minimum during January and February. All sample plots experienced a fall in water table during the summer months, although WT remained above -30 cm throughout the sample period. Both the lowest (-20 cm in 2014, -23 cm in 2015) and highest (0 cm in 2014, 2 cm above the surface in 2015) mean annual WT was observed in brash plots in both years. Over the course of the study, vegetation cover increased in the sample plots, particularly in the wetter Eriophorum vaginatum – Sphagnum plots. The C/N ratio in the soil profile (0-80 cm) in this site is 45.

159

250 (a) Year 1 200 Year 2 150

100

50 Precipitation (mm) Precipitation 0 March April May June July Aug Sept Oct Nov Dec Jan Feb

25 (b)

5 Soil temperature (°C)

0 Apr Jun Aug Oct Dec Feb Apr Jun Aug Oct Dec Feb 2014 2015 2016

Figure 6.2 Climate data for Scohaboy Bog, Co Tipperary during the sample period in 2014, 2015 and 2016. (a) monthly rainfall (mm) (Met Eireann, Gurteen Station).

Dark circles and line indicate 30 year average (1981-2010, www.met.ie), (b) soil temperature (°C) at 5 cm depth.

160

-2

-4

-6

-8 Water table (cm) table Water -10 Brash plots

-12 ------Sphagnum/ Eriophrium plots

-14 Cladonia/ Mosses plots -16 Apr May Jun Jul Aug Sep Oct Nov Dec Jan Feb Mar Apr May Jun

2014 Month 2015

Figure 6.3 Mean sample day water table in brash plots, Sphagnum/ Eriophorum plots and Cladonia/ Mosses plots from April 2014- July 2015.

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6.4.2 Measured methane fluxes

(a)

(b)

(c)

2014 2015

-2 -1 Figure 6.4 Observed CH4 fluxes (mg CH4 m h ) within the (a) Eriophorum vaginatum- Sphagnum, (b) Cladonia portentosa- Mosses and (c) Brash communities at Scohaboy Bog, Co. Tipperary. Positive values indicate a loss of CH4 to the atmosphere. Note differences in scale on the y-axis.

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Measured CH4 fluxes followed the sequence of brash > Cladonia portentosa- Mosses > Eriophorum vaginatum- Sphagnum plots (Figure 6.4). Brash plots had a mean water table depth of 7.6 cm. The variation in Eriophorum vaginatum- Sphagnum was less than other vegetation groupings. Methane emissions peaked in the summer months of July and August 2014 before falling steadily in the autumn months. Average sample day measured methane emissions varied from 0.105 mg -2 -1 -2 -1 CH4 m hr to 0.977 mg CH4 m hr . All sample plots acted as CH4 sources at all times for the duration of the study. The highest emissions were recorded on 6 August -2 -1 2014 in a brash plot (2.96 mg CH4 m hr ) while the lowest CH4 flux of 0.001 mg -2 -1 CH4 m hr occurred on 8 July 2014 in an Eriophorum vaginatum- Sphagnum plot. Emissions in the early summer months of 2014 were greater than those in 2015. A significant but weak relationship was found between the measured CH4 emissions and soil temperature (r2= 0.11, p< 0.05), a weak correlation was also found between 2 measured CH4 emissions and water table (r = 0.29, p< 0.05).

163

(a)

(b)

(c)

-2 -1 Figure 6.5 Measured N2O fluxes (μg m hr ) within the (a) Eriophorum vaginatum- Sphagnum, (b) Cladonia portentosa- Mosses and (c) Brash communities at

Scohaboy Bog, Co. Tipperary. Positive values indicate a loss of N2O to the atmosphere and negative values indicate N2O uptake by the peat. Note differences in scale on the y-axis.

Nitrous oxide emissions peaked in all plots in July during a period where WT across the study area dropped below 10 cm (Figure 6.5). From the end of July 2014 until

January 2015, N2O fluxes in most sample plots ceased or fell below the detection limit of the gas chromotograph. Water table depth and temperature both varied during this time period and so the trend in N2O emissions cannot be explained by any of the variables measured here. Nitrous oxide emissions became more variable

164 in spring as soil temperatures increased. Average sample day emissions ranged between -0.01 and 0.064 μg m-2hr-1.

6.4.3 Calculated Gas fluxes

2014 2015

-2 -1 Figure 6.6 Interpolated monthly means of CH4-C (g m month ) within the Eriophorum vaginatum- Sphagnum, Cladonia portentosa- Mosses and Brash communities at Scohaboy Bog, Co. Tipperary. Error bars represent standard deviation on the interpolated means.

Seasonal variation was evident and similar in all plots (Figure 6.6), following seasonal changes in temperature, (Figure 6.1) and WT (Figure 6.2). Fluxes tended to peak late in summer, coinciding with a rise in the WT. Spatial variation occurred across the sample plots, caused by varying WT, vegetation composition and the presence or absence of brash. Water table levels varied between 10 cm above and 28 cm below the peat surface in sample plots over the study period. Average sample period water levels ranged from -6 cm to -21 cm depth. Methane emissions in the course of this study did not strictly follow the water gradient, suggesting other

165 controlling factors take precedence in some plots. In general, brash plots tended to have higher emissions then vegetated plots but this difference was not significant. All sample plots in plough furrows maintained both similar water table levels and

CH4 emissions throughout the growing season. Sample plot vegetation displayed spatial variation across the site. Plots situated in the plough furrows were colonized by Sphagnum capillifolium, Sphagnum papillosum and Eriophorum vaginatum, those on the ridges consisted primarily of Cladonia portentosa, Sphagnum spp., Eriophorum vaginatum and Calluna vulgaris while brash plots remained bare.

Table 6.2 Summary of annual sums of CH4 and N2O and averages of relevant variables from the four study microsites. Standard errors of the estimates are shown in brackets for all fluxes. Positive values indicate a loss of C or N from the site.

T5 WT CH4-C N2O (g m-2 yr-1) (μg m-2 yr-1) Eriophorum- Sphagnum 13.5 -3.8 2.3 (0.03) 34.9 (25.1) Eriophorum 13.5 -4.6 2.25 43.6 Cladonia- Mosses 13.5 -4.6 2.4 (1.0) 180.6 (26.0) Brash 13.5 -7.5 4.8 (1.0) 29.8 (37.8)

T5 = soil temperature at 5cm depth in °C, WT = water table depth in cm. Soil temperature is common to all microsites as temperature was measured constantly at one location in the study site.

Interpolation of measured fluxes and subsequent averaging of annual balances of -2 -1 microsites (Table 6.2) led to an annual CH4 balance of 3.3± 0.6 g CH4-C m yr . High levels of variance are evident both temporally within and spatially between the sample plots. All plots were CH4 sources for the study period, with mean annual -2 -1 -2 interpolated emissions ranging from 1.4 g CH4-C m yr (S6) to 8.5 g CH4- C m -1 -2 -1 yr (S7). The estimated annual balance of N2O was 72 μg m yr . Scohaboy acted as a small source of N2O throughout the study period. Sample plots demonstrated great variation across the sample period during which the Cladonia portentosa- Mosses acted as the greatest source and the brash plots the smallest.

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6.5 Discussion

This is the first time that an investigation into the effect of rewetting on CH4 and

N2O emissions on rewetted raised peatland forests has been conducted in the temperate zone. The results presented here suggest that such sites remain low emitters of CH4 three years after rewetting. Estimated annual emissions ranging from -2 -2 2.3- 4.8 ± 1.0 g CH4-C m are lower than 6.5 g CH4-C m reported from rewetted - lawns in the Czech Republic for one growing season but higher than 1.0 g CH4-C m 2 for drier shrubby areas with a deeper WT in the same year (Urbanová et al., 2013). The mean WT measured in this study fluctuates between that of the lawns and shrubby areas in Urbanová et al. (2013). Studies from rewetted forests in Finland provide conflicting values of CH4 emissions. Losses from rewetted spruce swamps -2 -1 have been estimated as 14.5 g CH4-C m yr calculated from values reported by Koskinen et al., (2016) while Juottonen et al., (2012) recorded much smaller -2 -1 emissions of 0.082 – 2.96 g CH4-C m yr , also from rewetted peatland forest sites, highlighting the unpredictability of CH4 dynamics after rewetting. While the results of this study emphasise that rewetting recreates a CH4 source, it remains smaller than −2 −1 6.2 g CH4 m year , recorded on a natural Atlantic blanket bog in Kerry, southwest Ireland. Seasonal changes observed here were also seen in previous Irish studies on natural peatlands (Laine et al., 2007b). The pattern of fluxes peaking late in summer, coinciding with a rise in the WT, as has been found in another Eriophorum vaginatum dominated rewetted site (Marinier et al., 2004).

Previous studies have indicated that following rewetting CH4 emissions may initially remain low (Urbanová et al., 2012), due to the absence of labile organic matter on bare peat surfaces (Waddington and Day, 2007) and low methanogenic activity (Francez et al., 2000; Anderson et al., 2006). Although recently rewetted, and with patchy vegetation in many areas of the site, a large quantity of fresh organic material is available in Scohaboy due to the quantity of brash remaining. Brash collars emit -2 -1 the greatest volumes of all four microsite types at 4.8 ± 1.0 g CH4-C m yr .

Similarly, Mäkiranta et al., (2012) found that brash plots emitted more CH4 than control plots. It is expected that brash plots will continue as the largest CH4 emitters until the brash material is decomposed. As the other microsites increase in vegetation

167 coverage, emissions may increase due to increased volumes of organic material from senescing vegetation although input will likely remain less than from brash areas.

While initially methanogenic activity may be low in rewetted peatlands, rewetting encourages methanogen numbers to increase and development of the methanogenic community configuration towards that of natural peatlands (Urbanová et al., 2011), however, the numbers of both bacteria groups may remain lower than in natural sites

(Juottonen et al., 2012), limiting CH4 production. Therefore methane emissions from restored peatlands may still remain less than those of natural peatlands (Strack et al., 2013) as has been reported in this study.

Successful rewetting has been shown to initiate vegetation succession towards that of a natural peatland (Tuittila et al., 2000a; Komulainen et al., 2001). Vegetation change was observed on Scohaboy throughout the study period as Eriophorum vaginatum increased in both canopy size and spatial (more spread) extent. The occurrence of Sphagnum also increased in the sample plots. The plant species composition of a peatland can greatly influence CH4 dynamics (Robroek et al., 2015). In this study, larger fluxes occurred in the Cladonia- Mosses microsites than those dominated by Eriophorum vaginatum and Sphagnum spp. Aerenchymous plant species have been associated with higher CH4 emissions by a large body of research (e.g. Mariner et al., 2004; Leppälä et al. 2011, Green & Baird 2012), however, a smaller number of authors have connected reduced CH4 emissions from microsites containing aerenchymous species with an increased oxygenated rhizosphere (Mitsch and Gosselink, 2007; Dinsmore et al., 2009b). Sphagnum species have also been established as supporting methanotrophic bacteria (e.g. Raghoebarsing et al. 2005;

Larmola et al., 2009) and emitting lower volumes of CH4 emissions (Parmentier et al., 2011). Large fluxes in late summer can be attributed to a rising WT (Figure 7.2) and may also be related to the addition of decaying plant material to the peatland due to the onset of senescence (Tuittila et al., 2000b).

While important, vegetation control on CH4 emissions is less dominant than WT control (Leppälä et al., 2011). Water table levels were high in the Eriophorum vaginatum plots over the sample period, sometimes above the soil surface. High WT is associated with high CH4 emissions (Huttenun et al., 2003), however, in situations where the mean water levels are regularly above the peatland surface emissions may

168 be lower than those at or just below the surface (Couwenberg et al., 2011) negating the effect of aerenchymous species. Flooded conditions possibly constrain methane production by inhibiting CH4 diffusion through stationery water, while reduced vascular plant biomass contributes to a decrease in methane production (Turetsky et al., 2014). This phenomenon may have contributed to the lower emissions from the Eriophorum- Sphagnum plots.

The results of this study indicate that CH4 emissions increase following rewetting (Komulainen et al., 1998), however, three years following rewetting, they remain lower than those from natural sites. In the study area, Scohaboy bog retains a high water table throughout the year, fostering conditions for high CH4 emissions. As succession continues and microbe populations recover, provided WT depth remains shallow across the site, emissions may increase. Further long term investigation of rewetted peatland development in a temperate raised bog is required to track the evolution of the peatland and subsequent emission levels.

Nitrous oxide emissions are controlled by factors such as temperature, substrate, redox conditions, water table, pH and competition from vegetation (Machefert et al., 2002; Müller and Sherlock, 2004; Silvan et al., 2005; Lohila et al., 2010; Butterbach et al., 2013). Nitrous oxide emissions were low throughout the study period with no obvious temporal or spatial trend. Similar to Juszczak and Augustin (2013), N2O fluxes did not give a clear response to any particular variable. Emissions in some instances appear to be associated with WT drawdown, however, given the complexity of interactions governing N2O and the limited variables measured here, that cannot be confirmed. Our calculated annual balance of 72 μg m-2 yr-1, is much lower than 0.03- 0.92 g m-2 yr-1 reported from drained and forested peatlands in previous studies (Von Arnold et al., 2005; Ojanen et al., 2010). Rewetted peat extraction sites in the west of Ireland have detected no N2O emissions (Wilson et al.,

2013; Wilson et al., 2016b). In natural peatlands, N2O emissions have been assessed as being small or insignificant (Martikainen et al., 1993; Regina et al., 1996; Alm et al., 2007) although these may increase in peatlands drained for forestry (Minkkinen et al., 2002). Forest sites are fertilized when trees are planted and as fertilization contributes to higher N2O emissions (Augustin et al., 1998), residue of forestry fertilization may remain and be impacting emissions in this study, a situation absent from rewetted peat extraction sites. Other authors have concluded that N2O

169 emissions are insignificant from rewetted peatlands in temperate areas (Hendricks et al., 2007; Minke et al., 2015).

Given the size of emissions recorded, it is possible that small minus and positive values are errors around zero. In addition, due to the sampling method used in this study (stationary sampling points visited every two weeks), we may have missed ‘hot moments’ of emissions or ‘hot spots’ (Groffman et al., 2009; Butterbach et al., 2013) in the study site. As Scohaboy has a C/N ratio above 25, it was expected that

N2O fluxes would be small.

The water table for the duration of this study remained high creating water logged conditions. In long term saturated conditions, the redox potential decreases (Tauchnitz et al., 2015), influencing microbial activity in the soil, thereby reducing

N2O emissions (Rubol et al., 2012). Eriophorum vaginatum, which grows abundantly - on Scohaboy, has been identified as a strong competitor for NO3 , thereby moderating N2O emissions (Silvan et al., 2005). In other studies, logging residue on drained peatlands has contributed to an increase in N2O emissions (Mäkiranta et al., 2012). However, in this study, brash microsites did not display elevated emissions compared to the vegetated plots, most likely due to the raised WT. Rewetting was conducted three years prior to the study commencing, therefore due to the length of the hydrological recovery process on peatlands following drainage, it can be argued that the low levels measured in this study represent the peatland returning slowly to pre drainage conditions. The results of this study indicate that a newly rewetted raised peatland forest remains a small source of N2O three years post rewetting.

6.6 Conclusions

Long term monitoring on rewetted forestry sites is required in Ireland to trace their development as the effects of restoration influence vegetation development and gas dynamics. The restoration of a carbon sink function and pre-drainage hydrology following rewetting a long term process, however, has an effect on GHG emissions from peat almost immediately as seen in this project. Methane emissions from rewetted forestry return quickly to levels comparable with their pristine counterparts,

170 while N2O levels become almost negligible. Should the carbon sink function return to the site in time, it bodes well for the potential of the site to once again return to a natural functioning peatland.

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Chapter Seven

The impact of afforestation and subsequent rewetting on peat properties in blanket and raised bogs

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7.1 Abstract

Peat quality and composition impact on the rate of respiration in a peatland and subsequently CO2 and CH4 emissions. Soil organic matter (OM) decomposition is complex, comprising of physical, chemical and biological interactions. Lowering the WT through drainage exposes older, more recalcitrant peat to oxic conditions altering decomposition rates. Physical and chemical peat properties such as bulk density, ash content, pH and elemental composition are also modified by drainage and land use change. In this study, thermal analysis and elemental analysis of peat cores from natural, afforested and rewetted peatland sites was conducted to compare peat organic matter characteristics and chemical composition between the land use types. Physical properties; i.e. bulk density and ash content were also accessed between the three land uses. Pollagoona displays little variability in elemental composition between land use type, while significant differences were observed in C and N content between land uses in Scohaboy. Unexpectedly N content of the peat did not increase with depth in four of the six peat cores analysed and this is reflected in their C:N ratio. All cores contained both labile and recalcitrant OM. Refractory OM was found in forestry samples from both sites and Scohaboy natural. Comparisons between land use types indicate that drainage and subsequent rewetting alter the properties analysed in this study.

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7.2 Introduction

Previous chapters have outlined the changes peat undergoes following rewetting which influence GHG emissions. While gas analysis is vital to determine the climate change effect of drainage and rewetting on peatlands, determining and understanding the changes which soil properties and soil organic matter (OM) undergo following drainage for forestry and subsequent rewetting is essential to our understanding of and ability to alleviate the difficulties caused by both peatland drainage and rewetting.

Soil organic matter decomposition is complex, comprising of physical, chemical and biological interactions (Baldock et al., 2004). Soil organic matter decomposition can be characterized by three carbon pools, labile (fast), recalcitrant (slow) and refractory (inert) (Dell’ Abate et al., 2000; McLauchlan and Hobbie, 2004; Lopez- Capel et al., 2008). Carbon and N fluxes from soil are controlled by the highly reactive labile fraction, which in peat is influenced by WT, temperature and pH, while long term storage of C and N are dictated by the recalcitrant pool (Trumbore et al., 1990). Recalcitrance, the ability of OM to resist biodegradation, results from molecular-level characteristics of OM, including elemental composition, presence of functional groups and molecular conformation that influence their degradation by microbes and enzymes (Sollins et al., 1996).

Organic matter composition in peatlands varies with depth as a result of peat- forming conditions (Leifeld et al., 2012). Previous studies indicate that more labile components such as carbohydrate contents, carboxyl C and oxygenated reactions decline with depth while more thermally stable aliphatic C, aromatic C and lignin increase (Cocozza et al., 2003; Klavins et al., 2008; Delarue et al., 2011). Peat quality is greatly influenced by the growing vegetation composition which changes over time (Certini et al., 2015). Peat forming plants contain high levels of carbohydrates, primarily as cellulose that decomposes easily in soil (Hopkins et al., 1997). Some plants, especially mosses, such as Sphagnum, are a source of refractory compounds such as phenolics and aromatics (Williams and Yavitt, 2003). During peat formation, easily decomposable organic matter is lost and more recalcitrant and refractory material accumulates (Leifeld et al., 2012). It has been estimated that 80-

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90% of the net primary production deposited in the acrotelm of a peatland decomposes before moving into the anoxic, water saturated catotelm (Clymo, 1984). Turunen et al. (2002), estimates that the average rate of peat accumulation in natural Finnish peatlands is 18.5 g C m-2 yr-1. Drainage lowers the water table, increasing the oxic layer thereby exposing deeper more recalcitrant peat to oxic conditions, influencing the rate of peat decomposition (Similä et al., 2014). Peat quality and composition influence the respiration rate (Leifeld et al., 2012) and subsequently

CO2 and CH4 emissions (Reiche et al., 2010). Rewetting aims to return peatlands to C sinks, however in depth knowledge of the change in OM properties which also influence emissions after rewetting is lacking.

Soil organic matter has been widely studied using a range of methods providing chemical, physical and biological properties of SOM. Among these methods thermal analysis provides fast, easy procedures which supply data on the physical properties of OM (Barros et al., 2007 and Plante et al., 2009). These methods function by exposing samples to increasing temperature in order to detect several peaks in loss of weight (differential thermogravimetry; DTG) or energy release in the form of heat (differential scanning calorimetry; DSC). Two main peaks are shown in both DTG and DSC curves; a labile one at low or medium temperatures and a recalcitrant peak at higher temperatures. The oxidation of carbohydrates, proteins and disappearance of carboxyl groups causes the labile peak while polyphenolic and other structures including lignin produce the recalcitrant peak (Flaig et al., 1975 and Barros et al., 2007). Peat fractions are also represented by sample weight loss as relative proportions of the total weight loss (Exotot) as temperature increases, the first relates to the labile fraction (Exo1), the second to the recalcitrant (Exo2) and the third (Exo3) to the refractory fraction (Dell’ Abate et al., 2000; Dell’ Abate et al., 2002). To this author’s knowledge, no study has investigated the impact restoration has on peat OM using thermal analysis when compared to similar natural and drained forested peatland in a temperate region.

Drainage and land use change typically alters peat physical properties, i.e. bulk density, ash content and chemical properties such as pH and elemental composition. (Laiho and Laine, 1995; Laine et al., 1995; Minkkinen and Laine et al., 1998; Haapalehto et al., 2011; Leifeld et al., 2011a). Following rewetting, bulk density values have been found to remain high with small reductions in some instances

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(Tuittila et al., 2000a; Wilson et al., 2008; Zak et al., 2010) while the pH of peat extraction sites has risen (Lundin, 2012) until similar or higher than natural peatland (Wind-Mulder et al., 1996). A peatland’s C:N ratio is an indication of the level of peat decomposition (Malmer and Holm, 1984; Kuhry and Vitt, 1996). Larger C:N ratios are associated with slightly decomposed peat while the C:N ratio becomes smaller as peat becomes more decomposed owing to the loss of C during microbial decomposition. Simple methods for estimating C loss from drained peatlands have been developed using differences in ash content in a peat profile (Grønlund et al., 2008; Rogiers et al., 2008; Leifeld et al., 2011a). In a natural peatland, mineral matter is solely sourced from atmospheric deposition. These approaches are based on increased mineral matter (ash) accumulation following drainage and peat oxidation, i.e. increased loss of organic matter and the assumption that the ash content and bulk density in the catotelm is representative of undisturbed conditions. Methods outlined in the publications above use differences in bulk density and ash content to estimate C loss (Grønlund et al., 2008; Rogiers et al., 2008; Leifeld et al., 2011a; Leifeld et al., 2011b). In this study the bulk density, ash content, pH and elemental composition of a natural, forested peatland and rewetted blanket and raised peatland in Ireland were compared. We also carried out DTG and DSC analysis in order to assess the variety in thermal stability and quality of OM between natural, drained and rewetted peatland in Ireland. The aims of this work is to (a) compare the bulk density, ash content, pH and elemental composition of natural, forested and rewetted blanket and raised peatland in Ireland, (b) conduct thermal analysis (DTG and DSC) in order to assess the variety in thermal stability and quality of OM between the three treatments sites and (c) consider the implications of OM changes to C gas fluxes

7.3 Materials and Methods

7.3.1 Study Site

Soil properties were compared between natural, drained and forested and rewetted treatments in both Pollagoona and Scohaboy in order to identify differences between the treatments and potentially attribute them to land use. A description of the Pollagoona rewetted site is given in Section 3.3.1 and the Scohaboy site in Section

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3.1.1 In Pollagoona, cores from a natural area were taken from an adjoining natural site to the rewetted area. The drained forested sample site was 5.6 km away as a suitable site could not be found nearer to the rewetted area. Pollagoona natural site supports a wide diversity of plant species. Bog spicies found on the siteinclude Sphagnum spp, Calluna vulgaris, Deergrass (Scirpus cespitosus), Bog Asphodel (Narthecium ossifragum), Common Cottongrass (Eriophorum angustifolium) and the beak-sedges Rhynchospora alba and R. fusca. Bog-rosemary (Andromeda polifolia), has also been found on the site. Bog-myrtle (Myrica gale) is found at its western end of the bog. Lichens (Cladonia spp.) occur abundantly across the site. Pollagoona Bog contains slightly quaking flats of Bog Asphodel and beaksedges, with hummocks of Heather and mosses (Sphagnum spp. and Hypnum spp.). Due to its topographical location Pollagoona Bog does not appear to be adversely affected by the surrounding afforestation (NPWS, 2013b).

Natural site

Rewetted site

Figure 7.1 Rewetted and natural Pollagoona sites. Image was taken prior to tree removal from the rewetted site.

In Scohaboy, cores were taken from forestry and a natural peatland adjacent to the rewetted area. Scohaboy natural is made up of a large relatively flat raised bog with slopes associated with the nearby forestry. The bog supports vegetation typical of

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Midland Raised Bog type, consisting of Calluna vulgaris, Eriophorum spp., Bog Asphodel (Narthecium ossifragum) and White Beak-sedge (Rhyncospora alba). Other plants occurring on the site include Bog Rosemary (Andromeda polifolia) and Cranberry (Vaccinium oxycoccos). Sphagnum spp. cover is generally low with scattered tear pools in the north of the high bog which are beginning to fill with Sphagnum papillosum, S. capillifolium and S. cuspidatum. However, the scarce bog moss (S. imbricatum) is present in patches. In places in the north of the bog, lichen cover is high with abundant Cladonia portentosa (NPWS, 2013c).

Scohaboy Scohaboy Rewetted Forested

Scohaboy Natural

Figure 7.2 Scohaboy rewetted, natural and forested sites. Image was taken prior to tree removal from the rewetted site.

The proximity of both natural sites to the rewetted area may have caused them to be influenced by the hydrological changes implemented with drainage for forestry and subsequent rewetting. However, given that most peatlands have been influenced by anthropological activity and so an untouched peatland is difficult to find and their location exposes them to the same climatic conditions as the rewetted sites, these peatlands provided the best comparison of the natural state of each of these particular peatlands prior to drainage.

Following a visual assessment of the sites, considering vegetation and WT, sample points from which to take peat cores were selected. Four cores were taken to a depth of 80 cm using a box corer from each treatment site. Cores were split into 10 cm

181 segments on site, and refrigerated at 4 °C until prepared for air drying within 24 hours of sampling. pH was determined using fresh samples prior to air drying. Following air drying, samples, excluding those for pH analysis were sieved thorough a 2 mm sieve using a Rotar Mill (Fritsch P14 Variable Speed Rotar Mill, Gerhardt UK). Samples prepared for thermal analysis underwent no additional drying, samples for elemental analysis were further dried at 105°C until sample mass remained constant.

Physical, chemical and thermal properties are affected by land use and management in a peatland. In order to identify changes undergone by both peatlands following drainage for forestry and subsequent rewetting, carbon (C), nitrogen (N), sulphur (S), C;N ratio, ash content, bulk density and thermal properties of the peatland were examined.

7.3.1 Determining pH

pH was measured by preparing a solution of 10g of fresh soil in 25ml of distilled water, which was shaken briefly and allowed to stand for two hours before measuring the pH (Jenway 3310 pH meter).

7.3.2 Bulk Density

Following air drying, the whole sample was weighed to determine the mass of the air- dried soil. Two sub samples of approximately 5 g of the air dried soil were weighed out and placed into a brown paper bag individually. These were then dried in a Binder oven at 105°C for 24 hours. Oven dried samples were weighed to ±0.01g using an OHAUS Explorer scale. Using the soil moisture content of the 2 subsamples the water content of the whole sample was calculated. Equation (7.1) and -3 (7.2) below show how the soil moisture (%) and Db (g cm ) were calculated respectively.

푀푊− 푀퐴 SM = x 100 (Eq. 7.1) 푀푊 where:

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SM = soil moisture (%)

MW = Mass of the wet soil (g)

MA = Mass of the air dry soil (g)

푀퐴− 푀푂 Db = (Eq. 7.2) 푉푆 where: -3 Db = bulk density (g cm )

MA = Mass of the air dry soil (g)

MO = Mass of oven dry soil (g) 3 VS = Volume of soil (cm )

The air dried samples were then ground using a Fritsch pulverisette 14 rotary mill using a 2mm sieve after which samples were then stored in plastic bags or used for the analysis outlined below.

7.3.3 Elemental Analysis

Analysis of the carbon (C), hydrogen (H), nitrogen (N) and sulphur (S) content of soil was conducted. The C,H,N method of analysis adheres to ISO standard EN15104:2011, and S analysis follows the same procedure.

Prior to sampling, samples were prepared as follows:

1. Samples were removed from their containers and given time to equilibrate with the humidity of the air. 2. Moisture content of the sample was determined by overnight drying at 105°C, in duplicate. At least 1 g (measured to 0.1 mg) was weighed out in each crucible. Ash determination, as outlined in section 3.9.5 followed the next day using the oven-dried samples. 3. Approximately 50 mg of the sample was placed in a tin boat which was then wrapped up to prevent sample leakage. A 5 decimal place (i.e. 0.1 mg) analytical balance was used in this step. This was repeated a further 2 times

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for the same sample. The first replicate was used as a “run-in” in the sequence to condition the system for that sample and the analytical data was not used but instead was taken from the second and third replicates. 4. Standards (sulphanilamide) were prepared in the same way as for the samples. There was also a run-in followed by three standards. 5. Blank values were also determined using at least 3 analyses where no sample was processed. 6. The wrapped up samples were placed in the wheel of the vario MACRO cube and the sequence started using the following conditions: Combustion Tube Temperature at 1150°C and Reduction Tube Temperature at 850 °C.

Samples are placed on a rotating carousel at the top of the system and the software is programmed for their automatic analysis. The samples move to the furnace by means of a ball valve that rotates to receive the samples from the carousel. Outside air is removed by a helium or helium/oxygen gas mixture and the sample is then transferred to the combustion (C, H, N, S analysis) tube. Inside the combustion tube there is an ash crucible with a floor of aluminium oxide wool and it also contains differing compounds depending on particular analysis to be run. When set up for C, H, N, S, a tungsten oxide granulate sits below the ash crucible in order to deliver oxygen as a catalyst for combustion, prevent non-volatile sulphate formation and to bind alkali elements.

Helium acts as a carrier gas to transfer the gaseous products following combustion to the reduction tube. The tube is principally filled with copper which reduces the nitrogen oxides to nitrogen gas, sulphur trioxide is reduced to sulphur dioxide and any volatile halogens are bound to the silver wool also found in the tube.

Following this step, the products are passed to the separator columns. Nitrogen gas enters the thermal conductivity detector (TCD) first as it is not absorbed onto the columns. Carbon dioxide, H2O and SO2 are retained by their specific columns.

Following the detection of N2 on the TCD, each separator column in its turn is heated to its respective gases desorption temperature. This causes the gas that was adsorbed onto the column to be pushed along with the carrier gas until it ultimately reaches the TCD. The columns are then cooled for the next sample. Analysis of each replicate needs approximately 10 minutes.

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A peak for each element sampled will result from the TCD which is correlated to the quality of the element in the sample via calibrations. This requires determining blank values and daily factors for each of the elements. Blank values are measured by running an analysis with a C, H, N and S free sample and measuring the results. The blank values are subsequently subtracted from the corresponding elements of the real sample. Daily factors are used to fine tune the instrument calibration to the ambient conditions at analysis time and notice any differences in machine performance over time.

The elemental composition of C, N, H and S was calculated as a percentage by mass on a dry basis as follows;

100 C= Cad x ( ) (Eq. 7.3) 100−푀푎푑

100 N= Nad x ( ) (Eq. 7.4) 100−푀푎푑

푀푎푑 100 H= (Had- ) x ( ) (Eq. 7.5) 8937 100−푀푎푑

100 S= Sad x ( ) (Eq. 7.6) 100−푀푎푑

Where; ad is as determined;

Mad is the moisture content of the general analysis sample when analysed

7.3.4 Thermal Analysis

Thermogravimetry-differential scanning calorimetry (TG-DSC) was conducted using a Mettler Toledo TGA/DSC1 thermal analysis system. Soil samples were placed in 70μL aluminium oxide crucibles and heated. Temperature was increased by 10°C per minute from 50°C to 1000°C under a flux of 50 ml of air per minute. The organic matter content of the samples (OM) was determined as the weight loss between 200ºC-650 ºC and reported as a percentage on dry weight basis. Thermal analysis allows the differentiation of organic matter, Eox1 (labile organic matter) decomposes between 200°C-380°C; Exo2 (recalcitrant organic matter) decomposes between 380-

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475°C and Exo3 (refractory organic matter) accounts for weight losses between 475°C-650°C. The mass remaining after combustion is given as the percentage of char relative to the sample mass placed in the crucibles. Percentage of water in the sample is determined from the weight loss from 50 to 200 ºC following the DTG profile.

DSC curves are the profiles of the energy released by SOM during the combustion. DSC curves were integrated after correcting to a base line and the value of the integral was normalized to the OM content to give the heat of combustion of the samples (QSOM) in kJ g-1 OM. Recalcitrance (Exo1/Exotot) is given as a ratio of the three OM fractions in the peat. Increases in this ratio are indicative of increasing labile OM concentration and decreasing recalcitrance. TG-T50 and DSC-T50 were also determined for the samples. They are defined as the temperature at which 50% of the mass and energy respectively are lost and both are indices of thermal stability of matter.

7.3.5 Ash Content

Ash content was analysed following the ISO standard EN 14775: 2009. Empty crucibles were heated in an empty furnace at 550°C for 60 minutes. Crucibles were then removed from the furnace and allowed to cool for approximately 10 minutes before transferring to a desiccator where they remained until they reached ambient temperature. When cool, crucibles were weighed and the weight recorded to the nearest 0.1 mg. One gram of the peat sample was added to the crucible after which crucibles were weighed again and following further drying of the peat sample, the crucible and dish were reweighed in case of moisture absorption. Crucibles were then placed in a cold furnace and the furnace temperature raised according to the following procedure:

a) The furnace temperature was increased over a period of 30- 50 minutes to 250 °C at a heating rate of 4.5-7.5°C per min and subsequently maintained for one hour.

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b) The furnace temperature was raised to 550°C over 30 minutes (10°C per min) and maintained for two hours.

The crucible was removed from the furnace and allowed to cool for 5- 10 minutes before transferral to a desiccator. Once cool, the remaining ash and crucible was weighed to the nearest 0.1 mg. The ash content on a dry basis (Ad) of the peat sample is calculated as a percentage by mass on a dry basis, as per the formula below,

(푚3− 푚1) 100 Ad = x 100 x (Eq. 7.5) (푚2 − 푚1 ) 100−푀푎푑 where; m1 is the mass, in g, of the empty dish m2 is the mass, in g, of the dish plus the test sample; m3 is the mass, in g, of the dish plus ash;

Mad is the % moisture content of the test sample used for determination.

7.4 Results

7.4.1 Physical and Elemental Analysis of Natural, Forested and Rewetted Blanket and Raised Peat

Table 7.1 Summary of soil bulk density (g dm3) for Pollagoona and Scohaboy in each treatment

Depth Pollagoona Scohaboy (cm) Natural Forest Rewetted Natural Forest Rewetted 0-10 0.04 0.42 0.08 0.06 0.12 0.10 10-20 0.04 0.53 0.09 0.07 0.15 0.11 20-30 0.06 0.59 0.09 0.06 0.15 0.09 30-40 0.05 0.60 0.09 0.06 0.17 0.09 40-50 0.05 0.77 0.11 0.05 0.17 0.07 50-60 0.04 0.75 0.15 0.05 0.15 0.07 60-70 0.05 0.75 0.51 0.05 0.14 0.06 70-80 0.57 0.35 0.05 0.13 0.07 Mean 0.05 0.6 0.2 0.06 0.1 0.08

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Pollagoona natural ends at a depth of 60-70 cm as the site was so wet, deeper samples could not be taken.

Table 7.2 Ash content (%) per treatment in Pollagoona and Scohaboy in 10 cm increments

Depth Pollagoona Scohaboy (cm) Natural Forest Rewetted Natural Forestry Rewetted 0-10 4.45 14.55 4.82 4.99 4.2 3.76 10-20 3.61 36.48 2.73 3.19 3.25 2.43 20-30 2.61 47.59 2.51 1.84 3.41 1.69 30-40 2.64 30.06 2.47 1.74 3.35 1.58 40-50 2.75 17.55 3.01 1.5 5.33 1.53 50-60 3.47 13.67 12.82 1.67 7.37 1.71 60-70 4.32 16.37 54.1 1.53 9.15 1.58 70-80 16.4 78.97 1.55 9.92 1.53

Bulk density between both sites varies little with the exception of Pollagoona forest treatment (Table 7.1). Rewetted treatments lie between the two, suggesting rewetting has influenced the bulk density. In all treatments except Scohaboy natural and rewetted, there is an increase in bulk density with depth and mean bulk densities increase in the order of natural < rewetted < forest. Forest treatments had the greatest ash content per site in comparison to the other treatments (Table 7.2). An increase in ash content was observed in the Pollagoona forest site at 10-20, 20-30 and 30-40. A significant increase in ash content was also evident at 60-70 and 70-80 at the Pollagoona rewetted site. Scohaboy natural site displays a general decreasing trend in ash content while a general increasing trend in evident with depth in the Scohaboy forest site. The upper 20 cm of both natural treatments had higher ash content than the immediately lower layers, suggesting either an increase in atmospheric deposition or the influence of drainage. Ash content was significantly related to bulk density although the coefficient was relatively small (r2= 0.39, p<0.001).

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Table 7.3 Summary of soil pH for Pollagoona and Scohaboy in each treatment.

Depth Pollagoona Scohaboy (cm) Natural Forest Rewetted Natural Forest Rewetted 0-10 3.66 3.39 3.22 3.4 3.26 3.27 10-20 3.84 3.13 3.35 3.52 3.24 3.22 20-30 3.67 3.17 3.15 3.55 3.3 3.31 30-40 6.66 3.27 3.5 3.55 3.35 3.31 40-50 4.02 3.38 3.95 3.6 3.73 3.33 50-60 4.12 3.4 3.55 3.65 4.27 3.38 60-70 4.27 3.5 3.47 3.73 4.75 3.53 70-80 3.4 3.57 3.5 3.75 4.88 3.57

Table 7.4 Summary of soil carbon content (%) for Pollagoona and Scohaboy in each treatment.

Depth Pollagoona Scohaboy (cm) Natural Forest Rewetted Natural Forestry Rewetted 0-10 52.8 49.0 52.3 52.1 54.4 53.2 10-20 53.8 36.4 55.3 53.2 56.1 54.8 20-30 55.7 30.0 56.7 54.6 57.5 54.0 30-40 55.7 40.2 58.3 53.6 58.1 55.2 40-50 55.9 47.8 59.3 53.4 56.9 53.9 50-60 55.5 50.3 53.3 54.0 56.1 54.3 60-70 56.0 49.4 28.4 53.6 55.6 53.4 70-80 49.1 14.0 52.2 54.6 55.1

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Table 7.5 Summary of soil nitrogen content (%) for Pollagoona and Scohaboy in each treatment.

Depth Pollagoona Scohaboy (cm) Natural Forest Rewetted Natural Forest Rewetted 0-10 2.3 1.9 2.2 2.4 2.1 1.7 10-20 2.1 2.0 2.2 2.0 1.7 1.5 20-30 2.1 1.6 2.1 1.6 1.6 1.2 30-40 2.1 2.1 2.0 1.5 1.7 1.2 40-50 2.1 2.5 1.8 1.5 2.0 1.1 50-60 2.1 2.9 1.2 1.8 2.5 1.2 60-70 2.3 2.8 0.6 1.8 2.5 1.0 70-80 2.7 0.3 1.4 2.8 1.1

Table 7.6 Summary of C: N ratio for Pollagoona and Scohaboy in each treatment.

Depth Pollagoona Scohaboy (cm) Natural Forest Rewetted Natural Forest Rewetted 0-10 23.2 25.5 23.4 22.2 26.1 31.7 10-20 25.1 18.6 25.5 26.7 33.0 36.8 20-30 26.4 18.3 26.6 34.5 35.7 46.2 30-40 27.2 19.3 29.4 36.7 34.1 46.8 40-50 27.0 19.1 33.9 35.1 28.6 49.9 50-60 26.3 17.6 43.7 30.7 22.7 46.8 60-70 24.8 17.7 47.3 30.6 22.3 53.4 70-80 18.3 45.0 37.3 19.2 48.3

Table 7.7 Summary of soil sulphur content (%) for Pollagoona and Scohaboy in each treatment

Depth Pollagoona Scohaboy (cm) Natural Forest Rewetted Natural Forest Rewetted 0-10 0.50 0.32 0.35 0.36 0.28 0.32 10-20 0.57 0.37 0.41 0.46 0.28 0.44 20-30 0.64 0.30 0.33 0.46 0.29 0.39 30-40 0.69 0.43 0.30 0.39 0.30 0.36 40-50 0.71 0.54 0.22 0.37 0.43 0.42 50-60 0.73 0.63 0.20 0.36 0.64 0.37 60-70 0.84 0.58 0.10 0.39 0.81 0.33 70-80 0.72 0.07 0.35 0.91 0.36

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Soil characteristics tended to follow similar patterns in all treatments. At all treatments, the soil was acidic with mean pH values ranging from 3.4 (Scohaboy rewetted) to 4.2 (Pollagoona natural). As depth decreased down the peat profile, 0-80 cm, pH generally increased. Sulphur content generally increased across all treatments except for Pollagoona rewetted which decreased down the profile. In the upper layers of the soil profile except for the Pollagoona forestry, the C:N ratio was lower than in the deeper layers. As distance to the surface increased so too did the C: N ratio. The average C:N ratio was much lower in all Pollagoona treatments than in Scohaboy. The highest C:N ratio of the whole profile was found on both sites in the rewetted treatment and the lowest in the forestry treatment. One way ANOVA tests were used to analyse the differences between treatments. The effect of the treatments on soil non thermal soil properties in Pollagoona were evident only in nitrogen, sulphur content and the C:N ratio. Most elemental composition differences between treatments were non-significant (p> 0.05). Sulphur content varied significantly between all three Pollagoona treatments (p< 0.05) Rewetted samples showed a significantly higher (p= 0.028) C:N ratio than natural and forest samples (p< 0.05). A significant difference in pH was observed between the natural and forested (p= 0.03) treatments on Pollagoona, most likely due to the high pH observed at 30-40cm depth while all other differences in pH were non-significant. No significant difference in pH was found between any of the treatments on Scohaboy. Scohaboy showed significant differences in elemental composition between all treatments, except for sulphur content which was non- significant (p= 0.272). The changes observed in the Pollagoona soil data at depth where trends or characteristics are different between 60-80cm is most likely due to the increased influence of the mineral material beneath as Pollagoona is a shallow peatland.

191

7.4.2 Thermal Properties of Natural, Forested and Rewetted Blanket Peat

0.02

0.00

-0.02

-dm/dt -0.04

-0.06 (a)

-0.08 0 200 400 600 800 1000 Temperature (°C) 0.02

0.00

-0.02

-dm/dt -0.04

-0.06 (b) -0.08 0 200 400 600 800 1000 Temperature (°C) 0.02

0.00

-0.02

-0.04

-dm/dt -0.06

-0.08

-0.10 (c)

-0.12 0 200 400 600 800 1000 Temperature (°C) Figure 7.1 DTG curves of (a) natural, (b) forested and (c) rewetted blanket peatland at Pollagoona.

The first peak, 50-200°C is water and not considered an OM fraction. In the natural peatland, all depths display similar DTG characteristics, indicating that OM of all increments down the soil profile are constituted by both labile and recalcitrant fractions. The profile of the forested peatland differs, showing OM is made up of recalcitrant and labile fractions in the 0-10 layer, with a trend in the second peak to dissociate in two more fractions from 10-20 cm to 80 cm, indicating OM constituted by labile, recalcitrant and refractory fractions.

192

The rewetted sample is characterized by a change in properties as depth increases, starting with a bimodal behaviour in the upper layers, indicating OM is constituted by two fractions. The second peak tends to dissociate in a third peak at temperatures assigned to refractory material as depth increases. Therefore from 40-50 to 60-70 cm OM is constituted by three fractions. The third (Exo3) fraction disappeared in the deepest sample.

-5 OM

-1 -10

Wg -15 (a) -20 -25 0 200 400 600 800 1000 Temperature (°C) 2 0 -2 -4 -6 OM

-1 -8 -10 Wg -12 (b) -14 -16 -18 0 200 400 600 800 1000 Temperature (°C) 5

-5

OM -10 -1 -15 Wg -20 (c) -25

-30 0 200 400 600 800 1000 Temperature (°C)

Figure 7.2 DSC curves of (a) natural, (b) forested and (c) rewetted blanket peatland at Pollagoona.

193

The DSC curves of the natural peatland display the same characteristics down through the profile 0-80 cm and is constituted by both labile and recalcitrant fractions. The DSC curves of the forested peatland show changes in the OM nature of the profile as depth decreases as they evolve to increasing combustion temperatures. The rewetted sample indicates changes in the profiles of energy dissipation as depth increases, suggesting different chemical composition from upper to deepest layers.

(a)

120

100

20 OM thermal fractions thermal OM

0 0-10 10-20 20-30 30-40 40-50 50-60 60-70 (b) Depth (cm) 120

100

20 OM thermal fractions thermal OM

0 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 (c) Depth (cm) 120

100

20 OM thermal fractions thermal OM

0 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm) Figure7.3 Evolution of percentages of Exo1, Exo2 and Exo2 fractions with respect to total OM in (a) natural,(b) forested and (c) rewetted blanket peatland at Pollagoona.

194

The labile (Exo1) fraction of OM in the all three treatments tends to decrease with depth down the profile. In the natural peatland the recalcitrant fraction (Exo2) increases slightly with depth. In the forested sample, Exo1 fraction decreases with depth. The highest percentage of refractory OM (Exo3) is found in the 40-50 cm sample but generally, the refractory carbon increases with depth in the peat profile. The highest percentage of refractory material (Exo3) in the rewetted sample is at 40- 50 and 50-60 cm and tends to decrease and disappear in the 70-80 cm deepest sample. It is possible that the refractory fraction exists in the 10-20 to 30-40 increments but is overlapped with the recalcitrant fraction and cannot be quantified, in which case in the upper layers Exo 2 fraction is formed by recalcitrant and refractory material.

195

0.95

0.90 (a) 0.85

0.80

0.75

Exo1/Exotot Exo1/Exotot 0.70

0.65

0.60 0-10 10-20 20-30 30-40 40-50 50-60 60-70 Depth (cm) 0.80

0.75 (b)

0.70

0.65

0.60 Exo1/Exo tot Exo1/Exo

0.55

0.50 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm) 0.85 0.80 (c) 0.75 0.70 0.65 0.60 Exo1/Exotot 0.55 0.50 0.45 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm)

Figure 7.4 Recalcitrance levels in (a) natural (b) forested and (c) rewetted blanket peatland at Pollagoona.

In all three treatments OM recalcitrance is generally seen to increase as the ratio decreases. In the natural site, a decrease in recalcitrance is observed at 40-50 cm and this is reflected in Fig 8.3a as Exo1 (labile fraction) increases slightly. Similarly, the decrease in recalcitrance which occurs in the forested sample at 70-80 cm can be linked to an increase in Exo1 in Fig 8.3c.

196

18 (a)

OM) 17 -1

QSOM (kJ g (kJ QSOM 14

13 0-10 10-20 20-30 30-40 40-50 50-60 60-70 Depth (cm) 22 21 (b) 20 OM) -1 19

QSOM (kJ g (kJ QSOM 16

15 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm) 24

22 (c)

OM) 20 -1

QSOM (kJ g (kJ QSOM 14

12 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm)

Figure 7.5 Heat of combustion (QSOM) levels in (a) natural (b) forested and (c) rewetted blanket peatland at Pollagoona.

In the natural peatland, the OM chemical composition moved towards an increasingly oxidised state from 0-30 cm. These increments show surprising change in QSOM values, from those assigned to cellulose in the 0-10 increment, to a value lower than that for carbohydrates at 20-30 cm. In the lower layers, the OM remains stable at values associated with cellulose, increasing to that of lignocellulosic material in the 60-70 cm sample. Evidently the OM in this natural peatland is not greatly reduced and it is possible the changes in the upper layers are caused by higher microbial action there.

197

In the forested peatland, the QSOM values show that the OM in the 0-30cm increments is more reduced; being at the same degree of reduction as lignocellulosic material and lignin, then the deeper increments whose level of reduction is consistent with cellulose and carbohydrates. In the rewetted peatland, QSOM values shows an OM at a degree of reduction close to that of carbohydrates, showing a trend towards a slightly more oxidized state than carbohydrates as depth down the soil profile increases. It is very labile in terms of energy and that can explain the depletion with depth. This trend drastically changes towards high reduced OM in the deepest layers.

430 420 (a) 410 400 390 380 370 Temperature (°C) Temperature 360 350 0-10 10-20 20-30 30-40 40-50 50-60 60-70 Depth (cm) 480 460 (b) 440

420

400

380 Temperature (°C) Temperature 360 340 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm) 480

460 (c)

440

420

400

Temperature (°C) Temperature 380

360 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm) Figure 7.6 Thermal stability in (a) natural (b) forested and (c) rewetted blanket peatland at Pollagoona.

198

In the natural peatland, thermal stability is quite stable along the profile and the observed changes are small. The thermal stability of the organic matter on this site can be seen to increase to 30-40 cm and once again from 40-50 cm to 80 cm. In the forested peatland, the thermal stability of the OM in the 10-20 and 30-40 cm increments are similar after which it increases in terms of mass and energy content in the samples where the recalcitrant fraction decreases and the refractory fraction increases (40 to 60 cm). When the refractory material starts to decrease, thermal stability decreases too. The rewetted peatland sample shows that the material at 40- 50 cm is thermally stable in terms of OM loss, but below in lower increments this stability decreases and the sample is not thermally stable in terms of energy. This characteristic may cause OM to be lost rapidly despite its recalcitrance.

199

7.4.3 Thermal Properties of Natural, Forested and Rewetted Raised Peat

0.02

0.00 (a)

-0.02

-0.04 -dm/dt -0.06

-0.08

-0.10 0 200 400 600 800 1000 Temperature (°C) 0.02

0.00 (b)

-0.02

-0.04 -dm/dt -0.06

-0.08

-0.10 0 200 400 600 800 1000 Temperature (°C) 0.01 0.00 (c) -0.01

-0.02

-dm/dt -0.03

-0.04

-0.05

-0.06 0 200 400 600 800 1000 Temperature (°C)

Figure 7.7 DTG curves of (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy.

The first peak from 50 to 200 ºC is water and is not considered an OM fraction. The presence of refractory material is observed in the 0-20 cm and 50-60 cm increments, appearing as a shoulder in the DTG curve in Fig 7.7 (a). DTG curves of forested peatland show an OM made up primarily by two fractions, labile and recalcitrant. DTG curves evolve with depth indicating changes in the OM composition. From 30- 40 cm to 80 cm, the second peak shows a shoulder at about 490-520ºC indicating

200 refractory material, that cannot be quantified, is overlapped with the second DTG peak. The third peak at temperature higher than 600ºC can be attributed to mineral material. In the rewetted sample, DTG curves show a very similar behaviour with few changes with depth. Temperatures of the peaks indicate OM constituted by labile and recalcitrant fractions that could be attributed mainly to lignocellulosic material. These samples showed abundant un-degraded lignocellulosic material at every depth. Only samples from 0-20 cm formed ash after combustion. After that depth, all samples were constituted just by water and organic matter.

201

2 0 -2 -4 -6 OM

-1 -8 -10 Wg -12 -14 -16 -18 0 200 400 600 800 1000 Temperature (°C) 5

-5 OM

-1 -10

Wg -15

-20

-25 0 200 400 600 800 1000 Temperature (°C) 5

-5 OM -1 -10 Wg

-15

-20 0 200 400 600 800 1000 Temperature (°C) Figure 7.8 DSC curves of (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy.

In the natural peatland sample, the DSC curves evolve with depth, suggesting OM is made up of labile and recalcitrant fractions although in different ratios. Refractory material is observed in the 0-20 cm and 50-60 cm increments. Below 20 cm, the same OM properties are evident, while the increments above differ, implying that OM composition is slightly different above and below 20 cm. In the forestry sample, the DSC curves show changes in the OM nature of the profile from labile to recalcitrant, as depth decreases and they evolve to increasing combustion

202 temperatures. In the rewetted sample, DSC curves show the same OM properties from 0 to 20 cm and a slight change after that depth. (a)

120

100

OM thermal fractions thermal OM 20

0 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 (b) Depth (cm)

120

100

OM thermal fractions thermal OM 20

0 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 (c) Depth (cm) 120

100

OM thermal fractions thermal OM 20

0 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm)

Figure 7.9 Evolution of percentages of Exo1, Exo2 and Exo3 fractions with respect to total OM in (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy.

In the natural peatland, levels of labile (Exo1) and recalcitrant (Exo2) material remain relatively stable showing only slight increases and decreases throughout the profile except in the 60-70 increment where the quantity of Exo2 decreases and

203 refractory material (Exo3; poly condensed carbon, black carbon) is evident and then disappears in the last increment. In the forested peatland sample, (b), labile material tends to decrease with depth from 0-40 cm and to stabilize after, suggesting depletion of the microbial activity associated with labile SOM at depths higher than 40 cm. In (c) the proportion of labile and recalcitrant fractions (Exo1 and Exo2) from 0 to 20 cm are similar and Exo1 tends to increase slightly with depth after 20 cm, probably due to depletion of the microbial degradation of the labile material

(a)

(b)

(c)

Figure 7.10 Recalcitrance levels in (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy.

204

In the natural sample, recalcitrance trends vary down the profile, decreasing with depth from 0-50 cm and increasing from 50-70, where refractory material appears in the OM profile (Fig 7.9a), before decreasing again in the deepest increment. In sample (b) recalcitrance increases from 0 to 40 cm and it is caused by the depletion in the labile material. It depletes slightly after that depth and stabilizes. In sample (c), SOM recalcitrance decreases with depth due to the increasing Exo1.

(a)

(b)

(c)

Figure 7.11 Heat of combustion (QSOM) levels in (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy.

205

In the natural peatland profile QSOM shows clear changes in the nature of the SOM as depth increases. From 0-30 cm SOM has the same degree of reduction as carbohydrates. In the 30-50 cm increments, SOM increases in degree of reduction approaching that of cellulose and lignocellulosic material. Together with the observed evolution in Exo1/Exotot, that could indicate accumulation of that kind of material with depth maybe due to depletion of microbial degradation. SOM at 60-70 is the most reduced. In sample (b) the QSOM values indicate an evolution of OM to a high degree of reduction with depth. In the upper layers (0-40) it is at the same degree of reduction as lignocellulosic material, increasing after that to values indicating aromatic material from 40 to 70 cm. Both the natural and forested peatland samples show similar falls in reduction in the deepest increment to return to values indicative of lignocellulosic material. In sample (c) QSOM values tend to increase with depth but all values show OM at the same degree of reduction as that of carbohydrates.

206

480 460 (a) 440

420

400

380 Temperature (°C) Temperature 360

340 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm) 480

460

440

420

400

Temperature (°C) Temperature 380

360 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm) 460

440

420

400

380

Temperature (°C) Temperature 360

340 0-10 10-20 20-30 30-40 40-50 50-60 60-70 70-80 Depth (cm)

Figure 7.12 Thermal stability in (a) natural (b) forested and (c) rewetted raised peatland at Scohaboy.

In samples (a) the upper layers are the most thermally stable in the natural peatland sample and decreases with depth. Samples at 40-50 cm and 70-80 cm are the least stable and potentially will degrade the quickest. The forested peatland sample suggests that thermal stability of this peat increases with depth i.e. the OM at deeper layers would be more resistant to biodegradation than the upper layers. In the

207 rewetted sample, thermal stability generally decreases down the soil profile with a subsequent increase in T50 DSC in the lowest increment.

The contrasting behaviour of the peat down the profile with respect to DTG and DSC analysis is apparent by the variation in T50 for DTG and DSC (Fig 7.6 and Fig 7.12).

T50 values DTG are lower in all treatments on both sites than DSC. The behaviour of

T50 on Pollagoona tends to behave in a more erratic way than Scohaboy. Both T50 DTG and DSC in the forested treatments increase with depth indicating that there is an increase in the temperature required to release 50% of the energy stored and lose 50% of the peat OM in peatland forest soils, i.e. that forested peat is more thermally stable down the profile than either natural or rewetted peat. While the T50 DTG values of Pollagoona natural and rewetted treatments (Fig 7.6a and c) increase slightly with depth, they remains lower than their corresponding forested site. The

T50 DTG values of natural and rewetted samples from Scohaboy and T50 DSC from both sites decrease with depth, indicating decreasing thermal stability down the peatland profile in both natural and rewetted sites.

7.5 Discussion

7.5.1 Physical and Elemental Soil Properties

Estimates of the bulk density of northern peatland range from 0.068 g cm-3- 0.112 g cm-3 (Gorham, 1991; Turunen et al., 2002; Yu et al., 2006) while Connolly and Holden (2013) have estimated a bulk density of 0.10 g cm-3 for peatland in Ireland. Pollagoona natural falls just below this values at some depths (0.04 g cm-3) while Scohaboy is at the lower end of this range (0.05 g cm-3- 0.07 g cm-3). Drainage increases peat bulk density (Minkkinen and Laine, 1998), as evidenced by the greater values calculated for forestry sites in this study, although they remain lower than reported from some studies (Mustamo et al., 2016) while Scohaboy is similar to 0.095 ± 0.025 g cm-3- 0.146 ± 0.045 g cm-3 reported by Koskinen et al. (2016) from drained peatland forests in Finland. Rewetting of the peatland appears to have decreased the bulk density of the peatlands sampled, as rewetted bulk density values are lower than those calculated for forested sites although in Scohaboy the change is small in the 0- 20 cm increments. Pollagoona shows little difference in elemental composition between treatments. It is apparent from the results that although forestry

208 has affected both the nitrogen and sulphur content of the studied peatland, other elements and properties have not been influenced. Having been rewetted in 2006, Pollagoona had been influenced by a higher water table and a lack of trees for eight years prior to the study commencing, in comparison to Scohaboy, rewetted in 2013. Scohaboy shows much greater variability between treatments with significant differences seen between most elemental proportions. The elemental composition down the profile of most treatments across both sites is quite stable and most likely represents botanical composition and specific conditions when the peat was formed. The pH of forest and rewetted samples is lower than their natural counterparts. Following drainage, intensified oxidation of the organic material releases hydrogen ions into the peatland, increasing acidity (Laine et al., 1995). pH across all treatments measured in this study is lower than that measured in Ireland for natural peatland (Laine et al., 2007), drained agricultural peat (Renou- Wilson et al., 2014), rewetted industrial peatland (Wilson et al., 2016b) and similar or lower than peatland forest (Byrne et al., 2005). The similarity with values given in Byrne et al., (2005), as opposed to other peatlands in Ireland, is presumably caused by land use. Despite no active drainage occurring on our natural peatland, proximity to the rewetted sites may have impacted the ecosystem and therefore the properties sampled. Previous studies in Canada have reported an increase in sulphur concentration down a peat profile as was observed here (Moore et al., 2005).

Comparable C and N percentages have been detected in earlier studies (Urbanová et al., 2012; Renou- Wilson et al., 2014: Järveoja et al., 2016). Carbon content of most of the peat profiles examined in this study show an increase in the deeper layers, indicating that decomposition is increasing with depth. Pollagoona is a shallow peatland with mineral material beneath the peat and C content in the rewetted sample at 60 cm depth decreases as mineral material underneath the peat has a lower C content (Turunen et al., 1999). The percentage C in a peatland without mixing of the mineral material beneath increases with depth as C compounds are consumed and there is no inert material to increase in relative concentration (Tfaily et al., 2014). However, unexpectedly, N content did not increase with depth in most of the profiles which is reflected in the lack of consistency in C:N ratios in the soil profile. Nitrogen content in both forest treatments increased (excluding SF 0-10cm, the elevated 0- 10 cm increment most likely caused by fertilisation). The behaviour of C and N is

209 reflected in the evolution of OM fractions in many of the profiles (Fig 7.3 and Fig 7.9), where it is evident in PF, SN, SF and SR that the percentages of N content are reflected in the percentage changes of recalcitrant and refractory components.

The C:N ratios calculated in this study fall below those reported from earlier studies in natural peatland (Mustamo et al., 2016), rewetted peat extraction areas (Wilson et al., 2012; Järveoja et al., 2016) and forest sites are within the range reported for forested peatland in previous work (Ojanen et al., 2010). Ekono (1981) designates C values of 48-50% to slightly decomposed peat, 53-54% to moderately decomposed peat and 58-60% to highly decomposed peat. Only Pollagoona rewetted in the 30- 50cm zone and Scohaboy forestry in the 30-40 cm zone can be characterized as highly decomposed peat. The other profiles at the depths measured remain either slightly decomposed or moderately decomposed. Despite the substantial shift in peatland chemistry which occurs following drainage, the lack of significant differences between many of the soil properties in Pollagoona suggests that rewetting has successfully recreated natural peatland chemistry in relation to the properties which were sampled. Given the short time since rewetting in Scohaboy, it was expected that contrasts would be found between treatments. Despite this, bulk density values for the rewetted treatment are lower than forestry values, implying a shift in the peatland profile towards that of a natural peatland.

Peat depth in the Pollagoona area tends to be shallow, <1m in places, thereby causing some uncertainty in the justification of higher ash contents low down in the profile. While high ash values can often be attributed to peat drainage and other studies have utilized the ash residue method to make assumptions on time of drainage, subsidence and C loss since drainage (Rogiers et al., 2008; Leifeld et al., 2011b) we cannot be assured that the changes seen in our Pollagoona profile results are caused exclusively by drainage. It is likely that given the shallowness of peat surrounding the natural and rewetted site, there is some mineral soil from the parent material causing the increase seen from Pollagoona natural at 50-60 cm and Pollagoona rewetted at 50-60 cm depth. Ash peaks were also identified in Pollagoona forestry at 20-40 cm. Ash peaks represent periods of greater atmospheric deposition or other avenues of mineral deposition or episodes of increased peat decomposition. Ash peaks may also occur with changing peat accumulating rates, representing net peat loss or slowed peat growth (Leifeld et al., 2011a). The ash

210 residue method is certainly useful in deep peats and drained but not very disturbed peatland. In a shallow peatland, drainage ditches may cause the water table to drop too low to allow identification of background values not yet influenced by drainage (Rogiers et al., 2008). In Scohaboy, both in the rewetted and natural treatments there is a rise in ash concentration at the 10-20 cm increment which may be caused by the onset of drainage while the variability of ash concentration in the forest site highlights the level of disturbance and drainage that site has undergone. Deeper soil cores in drained, disturbed peatlands are needed to identify the background values of ash from the pre-drainage condition.

8.5.2 Thermal Soil Properties

Drainage, afforestation and rewetting cause changes in the thermal properties of OM on the peatlands studied. The first combustion peak in all DSC and DTG curves observed in this study occurs at temperatures normally assigned to labile substances, generally cellulose (Baffi et al., 2007) while reactions at temperatures above 400 °C are assigned to recalcitrant or refractory C, for example, lignin and aromatic compounds (Strezov et al., 2004; Dell ‘Abate et al., 2002). Pollagoona forestry and rewetted contains refractory material not found in the natural sample. As decomposition is inhibited in natural water logged conditions, it is expected that a higher percentage of labile material remains. In the drained environment, readily decomposable compounds decay, increasing the concentration of recalcitrant and refractory material in the peat profile as is evident from (Fig 7.3) with the appearance of Exo3, the refractory fraction in the forested and rewetted samples. Scohaboy shows the presence of refractory material in both the natural and forested samples of which only that in the natural sample could be quantified (Fig 7.9) and despite the percentage C indicating material is moderately decomposed at the lower depths (Fig 7.4). Recent vegetation on a site influences the composition of the C pools (Lopez- Capel et al., 2008) and so the vegetation growing in this area in the past is likely to be causing the occurrence of refractory material, particularly as the forested site if is the furthest away of the three treatment sites and may not have always had the same vegetation on site.

211

Recalcitrance is expected to decrease in natural peatland as occurs in Pollagoona, however, a greater level of recalcitrance is reached for the forestry and rewetted sample, indicating more decomposition of labile material. Recalcitrance in Scohaboy across the three profiles behaves differently to Pollagoona. In both the natural and rewetted treatments, recalcitrance decreases with depth and we assume age and the percentage of labile material increases, albeit in small quantities. Although not on peat soils, Kleber et al., (2010), reported an absence of recalcitrant, highly aromatic compounds in older C, suggesting that in some instances OM may not continue to decompose once it is formed. This therefore has consequences for our understanding of peat structure and storage capability. It is generally understood that stable OM makes up most of the C stored in soils (von Lützow and Kögel-Knabner, 2009), however, if in some instances, it is dominated by labile rather than recalcitrant material, other mechanisms must be preventing it from decomposing. In a peatland, we may attribute less decomposition in the peat to the water logged conditions seen in the natural and rewetted peatland, similar to that seen in (Reiche et al., 2010) and absent from the drained forested site. Heat of combustion values for OM further reveal how management or past vegetation composition may affect peatland. Despite all three peat treatments originating by the same process, all three have separate values down the profile. Furthermore, the combustion values also give us further information on the composition of the peat profiles as they are assigned to particular C fraction components.

The thermal stability of both sites is clearly affected by forestry, as thermal stability increases with depth on both peat types. Both the natural and rewetted sites behave in similarly to each other, although slightly differently between peat types. Natural and rewetted sits display a tendency to decrease in thermal stability with depth. Given the stark contrast in behaviour between forestry and other treatments, it appears that thermal properties of rewetted sites have returned close to those of their natural counterparts. Peat quality has been linked to CO2 and CH4 production rates in fens in Germany (Reiche et al., 2010), highlighting the need for further analysis using thermal analysis combined with gas measurement. Development of an OM quality index based on GHG emissions from natural, forested and rewetted peatlands could provide a method of estimating emissions rapidly from such sites.

212

7.6 Conclusion Drainage and subsequent rewetting has impacted on both physical and thermal properties of peat organic matter in this study. While many of the properties studied show a tendency to return to natural conditions, others such as heat of combustion values did not. Further study is required on peatland profiles using thermal analysis combined with GHG emission measurement to accurately link soil thermal properties and OM quality with emission levels.

213

214

Chapter Eight

General Discussion

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216

8.1 Greenhouse Gas balances

Greenhouse gas flux balances are dealt with in chapters’ four to seven. Chapter four outlines CO2 emissions from a rewetted blanket bog and chapter six details CH4 and

N2O emissions from the same site. Chapter five focuses on CO2 fluxes from rewetted raised bog while CH4 and N2O fluxes are discussed in chapter seven. The results of this study provide an assessment of greenhouse gas dynamics following rewetting. Despite the studied sites differing in peat type, they have similar dominant controls of GHG fluxes and study site characteristics. Pollagoona remains a source of CO2-C -2 -1 -2 (131.6± 298.26 g CO2-C m yr ) a small source of CH4- C (2.94±1.03 g CH4-C m -1 -2 -1 yr ) and a sink of N2O ( -11.78 μg m year ) eight years following rewetting however, one microsite, the Cladonia-Callluna plot was a modest C sink during the -2 -1 study year (-142.84 g CO2-C m yr ). Scohaboy Bog was a large source of CO2-C -2 -1 (585.3± 241.52 g CO2-C m yr ) and a small source of both CH4-C (3.25± 0.058 g -2 -1 -2 -1 CH4-C m yr ) and N2O (72 μg m year ), to the atmosphere during the monitoring year. It was concluded that WT is not the driver of CO2 fluxes at either site. Pollagoona displays a deep water table (ranging from -2 cm to -56 cm) in many plots over the course of the year and a very weak relationship was identified with WT but a much stronger one for temperature. Forest site preparation remains an influence in

Scohaboy. The highest RTOT values were recorded in the Sphagnum and Eriophorum micro-sites where a reasonably high mean annual WT of -4 cm was measured ranging from -21 cm in summer to +10 cm in winter while the smallest emissions were measured in the driest plots located on the ridges which are a legacy of the forest. However our results agreed with other studies (Lafleur et al., 2005; Renou- Wilson et al., 2014) who have previously demonstrated the control of soil temperature on RTOT. Soil temperatures recorded at both sites were relatively high and remained high throughout the winter period, maintaining high RTOT fluxes throughout the whole year as the mild winters in the temperate climate supports peat decomposition. High RTOT values in the furrows, despite shallow a WT, can be attributed to a greater layer of organic matter here than on the ridges (Saiz et al., 2006) which is likely to be decomposing even four years after felling. This release of C during the early rewetting stages is likely to last up to ten years (Höper et al.,

2008). Brash plots were the greatest sources of CO2-C as expected, confirming the

217 adverse effect of logging residue on peat GHG found in previous studies (Mäkiranta et al., 2012) and it is known that emissions can be sustained due to the priming effect (Fontaine et al., 2004) effecting emissions from Pollagoona even eight years post rewetting.

Despite no significant relationship being identified between CH4 and controlling variables, WT and temperature appeared to influence emissions. This is similar to other studies (Beetz et al., 2013; Renou-Wilson et al., 2013; Renou-Wilson et al.,

2016) that have used linear interpolation to calculate annual CH4 balances. Emissions in Pollagoona in particular follow the WT gradient and seasonal trends mimicking soil temperature are evident in both sites. Low emissions of both CH4 in this study are likely indicative of low quantities of methanogenic bacteria following drainage and subsequent rewetting (Juottonen et al., 2012) however as time since rewetting progresses emissions would be expected to increase (Waddington and Day, 2007), particularly when a high water table is maintained.

Vegetation is most significant in terms of C balance on both sites. The type of vegetation which colonizes a peatland following rewetting has a critical impact on the return of the C balance. Other authors have suggested vegetation as a dominant control on CO2 sequestration in the early stage of rewetting as the species of vegetation colonising the site and the amount of vegetation cover which a site displays can determine whether it is a sink or source (Samaritani et al., 2011; Urbanová, 2012). Pollagoona is fully vegetated eight years after rewetting. The site is been dominated by Molinia caerulea, with mosses such as Polytrichum commune, Hypnum cupressiforme and Sphagnum capillifolium present in smaller quantities, indicating the fluctuating water table and dry conditions of the site. Although Molinia caerulea has in previous work been found to be a C sink (Urbanová et al.,

2012), photosynthesis levels dropped while high amounts of CO2 continued to be released during senescence in this study causing a C source. In situations where other factors contribute to high C emissions, this may be detrimental to the C balance. Calluna, found in the plot which acted as a sink even in winter, is evergreen and can photosynthesise all year round unlike Molinia. Molinia also emitted the highest levels of CH4 during periods of warmer temperatures, similar to other studies (Ward et al., 2013). Scohaboy, as it is in the early stages of succession, lacks vegetation in some areas, particularly in the brash areas. Given the inability to model the fluxes

218 from the colonizing vegetation in the brash plots, it is likely that CO2 emissions were over estimated. During the course of the study, vegetation change was observed with Eriophorum vaginatum and Sphagnum species increasing in both spatial (more spread) and canopy (larger) extent. Eriophorum vaginatum influences C sequestration potential due to both its high respiration and productivity rates (Tuittila et al., 1999). It has also been found to be the largest C sink in rewetted cutaway blanket peatlands including at Bellacorick (Wilson et al., 2016b). Both species facilitate CH4 transport, (Carol and Freeman, 1999; Mariner et al., 2004), and also potentially influence microbes involved in CH4 dynamics (Robroek et al., 2015), thereby assisting in CH4 regulation. The vegetation dynamic on both sites in the future will have a significant impact on the C storage capacity of this site.

In addition to disrupting the C balance of a peatland, land use change typically alters peat physical properties (Laiho and Laine, 1995; Laine et al., 1995; Minkkinen and Laine et al., 1998; Tuittila et al., 2000a; Wilson et al., 2008; Zak et al., 2010; Haapalehto et al., 2011; Leifeld et al., 2011a; Lundin, 2012) as shown in the difference between properties of natural, drained and rewetted peatland sites sampled in this study. The thermal properties of the peatland treatments sampled here likewise changed in tandem with physical properties and hydrological conditions.

Strong relationships between CO2 emissions and peat organic chemistry have been well established (Turetsky et al., 2004; Reiche et al., 2010; Treat et al., 2014) and particular functional groups are indicative of potential emissions (Sjörersten et al., 2016). Thermal analysis indicates greater differences between Pollagoona natural and Pollagoona rewetted sites, than Scohaboy natural and rewetted sites. This may indicate that Scohaboy has greater potential to return to C sink function and natural conditions despite being a large C source soon after rewetting. Further analysis and monitoring is required to evaluate the success of thermal analysis as an indicator for C sink function return. This study highlights the need for an OM quality index based on GHG emissions from natural, forested and rewetted peatlands to provide a method of estimating emissions rapidly from such sites.

In this study, we did not consider losses of waterborne C. If included, the loss of C from the site could be even higher than reported here. For example, Renou-Wilson et al., 2014 reported 44 g C m−2 yr−1 and 30.8 g C m−2 yr−1 from grassland on organic soils in two consecutive years and Hendricks et al. (2007) found 20.6 ± 4.3 g C m−2

219 yr−1 was lost through water export. These examples highlight the proportion of fluvial losses of Cand therefore our C losses are likely underestimated.

8.2 Assessment of rewetting success

Rewetting is a long term approach to re-establish the C sink function of a peatland. It depends on a high WT and the vegetation present on the site (Drösler et al., 2008). Despite drain blocking, hydrological conditions have not been restored on Pollagoona as indicated by the low WT measured throughout the study year and the abundance of Molinia which is typically found on degraded blanket bog experiencing changeable WT levels. Therefore this study shows that a stable, high water table was not established following rewetting, preventing the return of the C sink function as the WT failed to reduce peat respiration and the colonisation of

Pollagoona by Molinia promoted CH4 emissions. However, CO2 emissions are greatly reduced compared to those from clear-felled peatland forests (Mäkiranta et al., 2007; Minkkinen et al., 2007) while CH4 emissions remain lower than in other rewetting projects, (Mariner et al., 2004: Wilson et al., 2013; Juottonen et al., 2012) and N2O emissions are negligible. The calculated GHG balance indicates that although the WT remains unstable, rewetting has effectively reduced emissions from the site although it still remains a source. Our results indicate that time since rewetting will reduce C emissions but abandonment after initial drain blocking and vegetation colonisation are not sufficient to return the C sink function in the absence of a stable high WT. Futher drain blocking needs to be carried out on Pollagoona in efforts to raise the WT further.

Hydrological conditions appear to have been reinstated in the study area of Scohaboy, although the site remains a large C source. Carbon dioxide losses from the site are comparable to and greater than some clear-felled peatland sites (Mäkiranta et al., 2007; Minkkinen et al., 2007) and peatland forests (Silvola et al.,

1996), however CH4 emissions continue at levels lower than a natural site in Co. Kerry (Laine et al., 2007b). The site is in the early stages of succession with slowly colonising vegetation, brash material left on site and a large volume of fresh organic matter available for decomposition and so this result is not surprising. Until peat forming vegetation has colonised the site and brash material has decomposed, this

220 site will continue to act as a drained organic soil releasing large volumes of CO2 despite a high water table being established.

Although both rewetting projects have failed to establish a C sink, Pollagoona and Scohaboy have potential for continued succession to C sinks. Current annual emissions from Pollagoona are lower than those from peatland forests and clear- felled peatland (Silvola et al., 1996; Mäkiranta et al., 2007; Minkkinen et al., 2007) and so rewetting has limited the OM loss from this site although it continues to act as a source. Scohaboy emissions remain similar to peatland forests and so rewetting has not instigated a change in gas dynamics in the short period since rewetting. Previous studies in Canada and Switzerland (Petrone et al., 2003; Samartitani et al., 2011) have proven that in some instances it takes many years for the natural conditions to be reinstated. However, should they fail to regain their C storage ability, other objectives such as structure or function of the peatland (Bonnett et al., 2009) could be assessed to determine levels of success. Other options to assess for success are hydrology, vegetation, birds, invertebrates and growth of peat (Wheeler and Shaw, 1995). Peatland supports a range of biodiversity (Peatland Ecology Research Group, 2009) and rewetting contributes to the re-establishment of valuable habitat.

This study has highlighted that rewetting peatland forests has not in the short term been successful in returning the C sink. Rewetted cutaway peatland in Ireland, has re- established a C sink after six years (Wilson et al., 2016) and maintains higher WTs than those recorded in this study. Greater collaboration between those conducting rewetting projects in Ireland is necessary to share techniques and develop methods for best practise. For future forest rewetting projects, it is recommended that all clear-felled trees be removed to remove the volume of fresh organic material on the site, limiting the C emissions from brash decomposition as both the results from this study and Mäkiranta et al. (2012) highlight that large quantities of CO2 are lost from peat underlying brash. Furthermore, drain blocking needs to be conducted to a greater extent and other work implemented to encourage a stable high WT thereby creating optimum conditions for peat forming vegetation colonisation. Research on drain- blocking methods from the U.K. (e.g. Armstrong et al., 2009) could be utilized to improve the effectiveness of Irish rewetting work.

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It is also recommended that sites be assessed carefully before commencing any tree- felling and rewetting. Changes in the peatland topography and hydrology following drainage will impact on the successfulness of rewetting and so careful attention needs to be paid to the hydrology of the site to determine how altered it is following drainage and will it be possible to re-establish and maintain a high, stable water table. Holden et al. (2011) highlighted that rewetting peatland sites may fail to re- establish natural water table dynamics even many years post rewetting. In situations where remedial works will not establish a high water table and the rewetted peatland will remain a source of C, it may in fact be more beneficial to leave the trees growing on the peatland. As forest drainage may result in a net gain of C depending on site and tree stand (Alm et al., 2007), as more C is stored in the growing trees is more than that lost from the drained peatland (Hargreaves et al., 2003), in situations where rewetting will not establish a C sink, the greater benefit may be from the C storage capacity of the trees.

8.3 Future Work

Future work should include continued monitoring of microsites in Scohaboy to track its successional development as time since rewetting increases. Insertion of collars into the natural peatland adjacent to the rewetted site will allow comparison between both conditions and further our understanding of the GHG controls on this site. Should further remedial works such as increased drain blocking be conducted on Pollagoona to raise the WT, continued monitoring would also track the changes that an older rewetted peatland undergoes to return to a C sink. Further thermal analysis of peat from all treatments, coupled with GHG sampling from each of the increments will enable us to understand the relationship between the OM at these sites and their GHG emissions and how land use change has affected them.

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Appendix 1

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Rigney, C., Wilson, W., Renou-Wilson, F., Müller, C. and Byrne, K. A. (2014) 'Methane emissions from rewetted peatland forests', in Science and Solutions for a Sustainable Environment Conference, University College Dublin, Ireland.

Methane emissions from rewetted peatland forests

Peatland management has undergone a number of policy changes in recent years. Globally, increasing trends are to rewet degraded peatlands with the overall aim of restoring the hydrology, ecosystem functions and biodiversity value characteristic of intact peatlands. Restoration of degraded peatlands is considered as a mitigation tool to combat climate change and the rising concentration of carbon in the atmosphere as rewetted and restored peatlands have the potential to act as large carbon stores in the future. In Ireland, Coillte has rewetted 2420.5ha of forest on blanket and raised peat. To date, little is known about how restored forest ecosystems function in terms of methane (CH4) dynamics. This project aims to measure CH4 fluxes in rewetted peatland forests and investigate the relationship of CH4 to environmental and climatic variables such as soil temperature, water table depth and vegetation type.

Methane emission factors will be derived from the results of this study. Two rewetted sites, previously under Coillte forestry, have been selected, one on blanket peat and the other on raised peat. Eight steel collars were inserted into each site to capture the range of topography across the site. CH4 fluxes are measured using the static chamber method. A 60 x 60 x 30cm chamber is inverted over the soil surface. Gas samples are drawn from the chamber headspace into evacuated glass bottles at 5 or 10 minute intervals and taken to the laboratory where the CH4 concentration is determined using gas chromatography. CH4 flux rates are calculated from the linear change of CH4 concentration in the chamber headspace as a function of time, base area, chamber volume and the molar volume of CH4 at chamber air temperature. Initial findings will be available for presentation at the meeting.

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Rigney, C., Wilson, W., Renou-Wilson, F., Müller, C. and Byrne, K. A. (2015) 'Methane Emissions from a Rewetted Blanket Peatland ', in Department of Life Sciences Postgraduate Research Day, University of Limerick, Ireland.

Methane Emissions in Rewetted Blanket and Raised Peatlands

In recent years the emphasis in peatland management has moved from productive functions towards restoration and rehabilitation. This is generally done by rewetting degraded peatlands with the overall aim of restoring the hydrology, ecosystem functions and biodiversity value characteristic of intact peatlands. Peatlands have the potential to impact significantly on climate change stemming from their capacity as a store or source of carbon. Restoration of degraded peatlands is considered a mitigation tool to combat climate change as it can restore the carbon sink function of these ecosystems. In Ireland, Coillte has rewetted 2420.5ha of formerly forested blanket and raised peat. To date, little is known about how these ecosystems function in terms of methane (CH4) dynamics. This project aims to measure CH4 fluxes in rewetted peatland forests and investigate the relationship of CH4 to environmental and climatic variables such as soil temperature, water table depth and vegetation type. CH4 emission factors will be derived from the results of this study. Two rewetted sites, previously under Coillte forestry, have been selected, one on blanket peat and the other on raised peat. Eight steel collars were inserted into the soil to capture the range of soil and vegetation across the site. CH4 fluxes are measured using the static chamber method. A 60 × 60 × 30 cm opaque chamber is inverted over the soil surface. Gas samples are drawn from the chamber headspace into evacuated glass bottles at 5 or 10 minute intervals and taken to the laboratory where the CH4 concentration is determined using gas chromatography. Methane flux rates are calculated from the linear change of CH4 concentration in the chamber headspace as a function of time, base area, chamber volume and the molar volume of CH4 at chamber air temperature. Initial findings will be available for presentation at the conference.

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Rigney, C., Wilson, W., Renou-Wilson, F., Müller, C. and Byrne, K. A. (2015) 'Methane Emissions in Rewetted Blanket and Raised Peatlands ', in Annual Assembly of the International Peat Society: ‘Peatlands a new conversation’, Tullamore, Co Offaly, Ireland.

Methane Emissions from a Rewetted Blanket Peatland In recent years the emphasis in peatland management has moved from productive functions towards restoration and rehabilitation. This is generally done by rewetting degraded peatlands with the overall aim of restoring the hydrology, ecosystem functions and biodiversity value characteristic of intact peatlands. Restoration of degraded peatlands is considered as a mitigation tool to combat climate change as it can restore the carbon sink function of these ecosystems. In Ireland, Coillte has rewetted 1212.3 ha of formerly forested blanket peat. To date, little is known about how these ecosystems function in terms of methane (CH4) dynamics. This project aims to measure CH4 fluxes in rewetted peatland forests and investigate the relationship of CH4 to environmental and climatic variables such as soil temperature, water table depth and vegetation type. CH4 emission factors will be derived from the results of this study. One rewetted peatland forestry site on blanket peat has been selected. Eight steel collars were inserted into the soil to capture the range of soil and vegetation across the site. CH4 fluxes are measured using the static chamber method. A 60 x 60 x 30 cm opaque chamber is inverted over the soil surface. Gas samples are drawn from the chamber headspace into evacuated glass bottles at 5 or

10 minute intervals and taken to the laboratory where the CH4 concentration is determined using gas chromatography. Methane flux rates are calculated from the linear change of CH4 concentration in the chamber headspace as a function of time, base area, chamber volume and the molar volume of CH4 at chamber air temperature. Initial findings will be available for presentation at the conference.

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Rigney, C., Wilson, W., Renou-Wilson, F., Müller, C. and Byrne, K. A. (2016) 'Carbon dioxide dynamics of a rewetted peatland forest on Blanket Peatland ', in Environ 2016, University of Limerick, Ireland.

Carbon dioxide dynamics of a rewetted peatland forest on Blanket Peatland

In recent years the emphasis in peatland management has moved from productive functions towards restoration and rehabilitation. This is generally done by rewetting degraded peatlands with the overall aim of restoring the hydrology, ecosystem functions and biodiversity values characteristic of intact peatlands. Restoration of degraded peatlands is considered as a mitigation tool to combat climate change as it can restore the carbon sink function of these ecosystems. In Ireland, Coillte has rewetted 1212.3 ha of formerly forested blanket peat. To date, little is known about how these ecosystems function in terms of carbon dioxide (CO2) dynamics. This project aims to measure carbon dioxide (CO2) fluxes in a rewetted peatland forest on blanket bog and investigate their relationship to environmental and climatic variables such as soil temperature, water table depth and vegetation type. CO2 emission factors will be derived from the results of this study. Eight steel collars were inserted into the soil and CO2 fluxes were measured using a 60 × 60 × 30 cm static polycarbonate chamber, fitted with a battery operated fan and cooling apparatus, which is inverted over the soil surface. Carbon dioxide concentration (ppm) in the chamber headspace is measured with a portable analyser and CO2 fluxes are calculated from the linear change of CO2 concentration in the chamber headspace as a function of time, temperature and chamber volume. Results will be available for presentation at the conference.

Keywords: Carbon dioxide, blanket peatland, rewetted, restoration, deforested

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Rigney, C., Wilson, W., Renou-Wilson, F., Müller, C. and Byrne, K. A. (2016) 'Carbon dioxide dynamics of a rewetted peatland forest on Blanket Peatland', in Department of Life Sciences Postgraduate Research Day University of Limerick, Ireland.

Carbon dioxide dynamics of a rewetted peatland forest on Blanket Peatland

(ppm) in the chamber headspace is measured with a portable analyser and CO2 fluxes are calculated from the linear change of CO2 concentration in the chamber headspace as a function of time, temperature and chamber volume. The study site was a source -2 -1 of CO2 with an annual carbon balance of 154.5g CO2 – C m yr for year one of the study. These results suggest that rewetting peatlands reduces carbon emissions and confirms that returning the carbon balance takes many years.

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Rigney, C., Wilson, W., Renou-Wilson, F., Müller, C. and Byrne, K. A. (2016) 'Carbon dioxide exchange in a rewetted peatland forest on blanket peatland', in Inaugural International Conference on Natural and Constructed Wetlands, National University of Ireland, Galway, Ireland,

Carbon dioxide exchange in a rewetted peatland forest on blanket peatland

Introduction and Objectives In recent years the emphasis in peatland management has moved from productive functions towards restoration and rehabilitation. This is generally done by rewetting degraded peatlands with the overall aim of restoring the hydrology, ecosystem functions and biodiversity values characteristic of intact peatlands. Restoration of degraded peatlands is considered as a mitigation tool to combat climate change as it can restore the carbon sink function of these ecosystems. In Ireland, Coillte has rewetted 1212.3 ha of formerly forested blanket peat. To date, little is known about how these ecosystems function in terms of carbon dioxide (CO2) exchange. This project aims to measure CO2 fluxes in a rewetted peatland forest on blanket bog and investigate their relationship to environmental and climatic variables such as soil temperature, water table depth, photosynthetic photon flux density (PPFD) and vegetation type. Carbon dioxide emission factors will be derived from the results of this study.

Materials and Methods Pollagoona is a blanket bog located in north County Clare. The restoration site totalled 60.6 ha of which 40.3 ha was forested; predominately with Sitka Spruce. The main restoration activities which took place were manual felling of conifers, windrowing in places and the blocking of manmade drains. Eight steel collars were inserted into the soil prior to commencing gas exchange measurements. Light and dark chambers were used to measure instantaneous net ecosystem exchange (NEE). NEE was measured using a static polycarbonate chamber (60 cm × 60 cm × 30 cm) fitted with a battery operated fan, whose function was to mix the air within the chamber headspace. The chamber was connected to a cooling system which maintained the internal temperature to within 1ºC of the ambient temperature. CO2 fluxes were measured using a portable infrared gas analyser (EGM- 4, supplied by

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PP Systems). Carbon dioxide concentration (ppm) in the chamber headspace was measured with a portable analyser and CO2 fluxes are calculated from the linear change of CO2 concentration in the chamber headspace as a function of time, temperature and chamber volume. CO2 fluxes were measured in both the light and the dark in order to determine total ecosystem (heterotrophic and autotrophic) respiration (RTOT) (Method as per Wilson et al 2007a, Tuittila et al., 1999 and Tuittila et al., 2004). Simultaneously, air temperature and PPFD was recorded. Carbon flux measurements were conducted on a biweekly or monthly basis in the period August 2014 until September 2015.

Results and Discussion 2 RTOT displays a strong relationship, (r = 0.74) with temperature at 5 cm at this site.

The models used for RTOT, based upon the Arrhenius equation (Lloyd and Taylor, 1994), are non-linear models relating to soil temperature. The influence of temperature on peat respiration has been well documented (Riutta et al., 2007a; Wilson et al., 2007b). A strong relationship was observed between GPP and PPFD at this site, where it accounted for 55% of the variation. The addition of temperature increased the explanatory power to 79% at some collars while the addition of green area index resulted in 77% of the variation being explained in other collars. Other studies have also modelled temperature to explain variability in GPP (Riutta et al., 2007b; Laine et al., 2009; Wilson et al., 2015). Water table appears to have little or no control on GPP or RTOT at this site. It may be that fluctuations in WT level were missed due to the flux measurement regime employed in this study. As water table remained below 10 cm depth at many of the collars during the duration of this study, it may be that our results reflect the complexity of the relationships between variables in dry soils. Acknowledgements

This project is funded by the Environmental Protection Agency (2012-CCRP- PhD. 2). Grateful thanks to Coillte for access to their LIFE 02 and LIFE 09 sites. Many thanks to Michael Kenny (Carlow Institute of Technology), Michael Kenna (Noone Engineering, Rathangan, Co Kildare) and Ray Byrne for their assistance with equipment design and manufacture.

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References

Laine, A., Riutta, T., Juutinen, S., Vӓliranta, M., and Tuittila, E. S., 2009, Ecological Modelling, 2646- 2655

Lloyd, J. and Taylor, J. A., 1994, Functional Ecology, 315–323

Riutta, T., Laine, J., Aurela, M., Rinne, J., Vesala, T., Laurila, T., Haapanala, S., Pihlatie, M, and Tuittial, E.S., 2007, Tellus, 838-852

Riutta, T., Laine, J., and Tuittia, E.S., 2007, Ecosystems, 718-733

Tuittila, E.S., Komulainen V-M., and Vasander H., 1999, Oecologia, 563- 574

Tuittila, E.S., Vasander, H., and Laine, J., 2004, Restoration Ecology, 483-493

Wilson, D., Alm, J., Riutta , T., Laine, J., Byrne, K. A., Farrell, E. P. and Tuittila, E. S., 2007, Plant Ecology, 37- 51.

Wilson, D., Tuittila, E. S., Alm, J., Farrell, E. P., Byrne, K. A., 2007, Ecoscience, 71- 80

Wilson, D., Dixon, S. D., Artz, R.R.E, Smith, T.E.L., Evans, C.D., Owen, H.J.F., Archer, E., and Renou- Wilson, F., 2015, Biogeosciences, 5291- 5308

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