Volatile Organic Compounds in the Atmosphere

Volatile Organic Compounds in the Atmosphere

Volatile Organic Compounds in the Atmosphere Edited by Ralf Koppmann University of Wuppertal, Germany

Blackwell Publishing

© 2007 by Blackwell Publishing Ltd Blackwell Publishing editorial offices: Blackwell Publishing Ltd, 9600 Garsington Road, Oxford OX4 2DQ, UK Tel: +44 (0)1865 776868 Blackwell Publishing Professional, 2121 State Avenue, Ames, Iowa 50014-8300, USA Tel: +1 515 292 0140 Blackwell Publishing Asia Pty Ltd, 550 Swanston Street, Carlton, Victoria 3053, Australia Tel: +61 (0)3 8359 1011 The right of the Author to be identified as the Author of this Work has been asserted in accordance with the Copyright, Designs and Patents Act 1988. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system, or transmitted, in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, except as permitted by the UK Copyright, Designs, and Patents Act 1988, without the prior permission of the publisher. First published 2007 by Blackwell Publishing Ltd ISBN: 978-1-4051-3115-5 Library of Congress Cataloging-in-Publication Data Volatile organic compounds in the atmosphere/edited by Ralf Koppmann. – 1st ed. p. cm. Includes bibliographical references and index. ISBN-13: 978-1-4051-3115-5 (hardback : alk. paper) 1. Organic compounds–Environmental aspects. 2. Air quality management. I. Koppmann, Ralf. TD885.5.O74V66 2007 511.51 12–dc22 2006034260 A catalogue record for this title is available from the British Library. Set in 10/12 pt Minion by Newgen Imaging Systems (P) Ltd., Chennai, India Printed and bound in Singapore by Markono Print Media Pte Ltd The publisher’s policy is to use permanent paper from mills that operate a sustainable forestry policy, and which has been manufactured from pulp processed using acid-free and elementary chlorine-free practices. Furthermore, the publisher ensures that the text paper and cover board used have met acceptable environmental accreditation standards. For further information on Blackwell Publishing, visit our website: www.blackwellpublishing.com

Contents

Preface

ix

List of Contributors

xi

1

1

2

3

Volatile Organic Compounds in the Atmosphere: An Overview Jonathan Williams and Ralf Koppmann 1.1 Introduction 1.2 Sources 1.3 Sinks 1.4 Atmospheric distribution 1.5 Measurement tools 1.6 Modelling tools 1.7 How organic species affect the atmosphere 1.8 Open questions and future directions References

1 3 5 7 9 10 12 15 19

Anthropogenic VOCs Stefan Reimann and Alastair C. Lewis 2.1 Introduction 2.2 Sources of anthropogenic VOCs 2.3 Atmospheric distribution of VOCs 2.4 Chemical behaviour of VOCs in the atmosphere 2.5 Measurement techniques References

33

Biogenic VOCs Allison H. Steiner and Allen L. Goldstein 3.1 Introduction 3.2 Sources of biogenic VOCs 3.3 Emission inventories of biogenic VOCs 3.4 Global distribution of biogenic VOCs 3.5 Impact on photooxidants and atmospheric chemistry

82

33 33 45 55 60 70

82 83 97 103 107

vi

4

5

6

7

Contents

3.6 Sampling and measurement techniques 3.7 Future directions References

114 116 117

Oxygenated Volatile Organic Compounds Ralf Koppmann and Jürgen Wildt 4.1 Introduction 4.2 Tropospheric mixing ratios and global distribution 4.3 Sources of OVOCs 4.4 Sinks of OVOCs 4.5 Budgets and emission inventories 4.6 Sampling and measurement techniques 4.7 Future directions Acknowledgement References

129

Halogenated Volatile Organic Compounds Simon J. O’Doherty and Lucy J. Carpenter 5.1 Introduction 5.2 Sources of halogenated VOCs 5.3 Atmospheric concentrations: trends and distribution 5.4 Sinks of halogenated VOCs 5.5 Emission inventories 5.6 Sampling techniques 5.7 Measurement techniques References

173

PAN and Related Compounds James M. Roberts 6.1 The chemistry of PANs 6.2 Atmospheric formation 6.3 Measurement and calibration techniques 6.4 Atmospheric measurements 6.5 Modelling and interpretation of ambient measurements 6.6 Conclusions Acknowledgements References

221

Organic Nitrates Paul B. Shepson 7.1 Introduction 7.2 Production mechanism 7.3 Measurement methods 7.4 Atmospheric measurements 7.5 Fate 7.6 Conclusions References

269

129 130 137 149 154 155 160 160 160

173 179 187 192 204 207 210 214

222 229 237 243 249 255 256 256

269 271 274 276 282 285 286

Contents

8

9

vii

High-Molecular-Weight Carbonyls and Carboxylic Acids Paolo Ciccioli and Michela Mannozzi 8.1 Introduction 8.2 Sources 8.3 Atmospheric levels 8.4 Reactivity and impact on the atmosphere 8.5 Sampling and analysis 8.6 Conclusions References

292

Organic Aerosols Thorsten Hoffmann and Jörg Warnke 9.1 Introduction 9.2 Carbonaceous aerosols 9.3 Analysis of organic aerosols Further reading References

342

292 293 309 324 329 333 334

342 345 365 375 375

10 Gas Chromatography-Isotope Ratio Mass Spectrometry Jochen Rudolph 10.1 Introduction 10.2 Fundamentals of stable isotope ratios of VOCs 10.3 Experimental methods 10.4 Kinetic isotope effects 10.5 Stable isotope ratios of atmospheric VOC and their sources 10.6 Conclusions References

388

11 Comprehensive Two-Dimensional Gas Chromatography Jacqueline F. Hamilton and Alastair C. Lewis 11.1 Introduction 11.2 Fundamentals of comprehensive gas chromatography 11.3 Modulators 11.4 Detectors 11.5 Examples of GC × GC use in atmospheric samples 11.6 Conclusions Further reading References

467

Index

489

Color plate appears between pages 268 and 269

388 389 405 420 447 458 460

467 468 471 474 475 482 486 486

Preface

Every day, large quantities of volatile organic compounds (VOCs) are emitted into the atmosphere from both anthropogenic and natural sources. They are the ‘fuel’ that keeps atmospheric photochemistry running. Therefore, their sources, sinks and residence times are the subject of current research. In addition to influencing local, regional and even global photochemistry, several of these compounds have a potential impact on climate, both due to their properties as greenhouse gases and due to their ability to form aerosol particles on oxidation. The formation of gaseous and particulate secondary products caused by the oxidation of VOCs is one of the largest unknowns in the quantitative prediction of the earth’s climate on a regional and global scale, and on the understanding of local air quality. To be able to model and control their impact, it is essential to understand the sources of VOC, their distribution in the atmosphere and the chemical transformations they undergo. Furthermore, organic trace gases can be used as tracer compounds to investigate reactions that are not directly accessible to current measurement techniques or as probes to ‘visualise’ transport processes in the atmosphere or across atmospheric boundaries. In recent years methods and techniques for the analysis of organic compounds in the atmosphere have been developed to increase both the spectrum of detectable compounds as well as the corresponding detection limits. New methods have been introduced to increase the time resolution of those measurements and to resolve more complex mixtures of organic compounds. New technical developments reducing weight and energy requirements made the use of these instruments on various platforms such as balloons or aircraft possible. This book describes the current state of knowledge of the chemistry of VOC as well as the methods and techniques to analyse gaseous and particulate organic compounds in the atmosphere. Chapter 1 is an instructive chapter summarising the variety and the roles of VOC in the atmosphere. Chapters 2 to 9 cover the various compound classes, their distribution in the atmosphere, their chemical transformations and their budgets as well as a survey of currently used measurement techniques. Chapters 10 and 11 describe new methods to measure a large part of the VOC family at a glance and for investigating their stable carbon isotope ratios. In-depth references are provided, enabling each subject to be explored in more detail. The aim is to provide an authoritative review to address the needs of both graduate students and active researchers in the field of atmospheric chemistry research. It may also serve as a desktop resource for experienced scientists in the field of atmospheric research.

x

Preface

Thanks are due to all the chapter authors for their efforts in the completion of this work. I am grateful to many colleagues for numerous discussions, for their patience, their advice and critical reviews of the chapters. Special thanks are due to Sarahjayne Sierra and Dr Paul Sayer of Blackwell Publishing for their patience in answering all my questions and their persistence in the efforts required for completing this book.

List of Contributors

Dr Lucy J. Carpenter

Department of Chemistry, University of York, York, UK

Dr Paolo Ciccioli

Istituto di Metodologie Chimiche del C.N.R., Monterotondo Scalo, Italy

Professor Allen L. Goldstein

Department of Environmental Science, Policy, and Management, University of California, Berkeley, CA, USA

Dr Jacqueline F. Hamilton

Department of Chemistry, University of York, York, UK

Professor Thorsten Hoffman

Institut für Anorganische und Analytische Chemie, Johannes Gutenberg-Universität Mainz, Mainz, Germany

Professor Ralf Koppmann

Fachbereich Mathematik und Naturwissenschaften, Atmosphärenphysik, Bergische Universität Wuppertal, Wuppertal, Germany

Professor Alastair C. Lewis

Department of Chemistry, University of York, York, UK

Dr Michela Mannozzi

Istituto Centrale per la Ricerca Scientifica e Tecnologica Applicata al Mare, Roma, Italy

Dr Simon J. O’Doherty

School of Chemistry, University of Bristol, Bristol, UK

Dr Stefan Reimann

Eidgenössische Materialprüfungs- und forschungsanstalt, EMPA, Duebendorf, Switzerland

Dr James M. Roberts

NOAA Earth Systems Research Laboratory, Chemical Science Division, Boulder, CO, USA

Professor Jochen Rudolph

Chemistry Department and Centre for Atmospheric Chemistry, York University, Toronto, Ontario, Canada

Professor Paul B. Shepson

Departments of Chemistry, and Earth and Atmospheric Sciences, Purdue University, West Lafayette, IN, USA

Professor Allison H. Steiner

Department of Atmospheric, Oceanic and Space Science, University of Michigan, Ann Arbor, MI, USA

xii

List of Contributors

Dr Jörg Warnke

Institut für Anorganische und Analytische Chemie, Johannes Gutenberg-Universität Mainz, Mainz, Germany

Dr Jürgen Wildt

Institut für Chemie und Dynamik der Geosphäre, Institut 3: Phytosphäre, Forschungszentrum Jülich, Jülich, Germany

Dr Jonathon Williams

Atmospheric Chemistry Department, Max Planck Institute for Chemistry, Mainz, Germany

Volatile Organic Compounds in the Atmosphere Edited by Ralf Koppmann Copyright © 2007 by Blackwell Publishing Ltd

Chapter 1

Volatile Organic Compounds in the Atmosphere: An Overview Jonathan Williams and Ralf Koppmann

1.1

Introduction

The aim of this overview is to highlight the importance of organic trace gases in the atmosphere and to introduce the themes of the chapters to follow. This work is suited to those new to the field and to those seeking to place related activities in a broader context. Tens of thousands of organic compounds have been detected in the air we breathe, and the focus here is on the myriad carbon-containing gases present at mixing ratios of some 10 parts per billion (ppbv, 10−9 or nmol/mol) down to some parts per trillion (ppt, 10−12 or pmol/mol). This excludes the three most abundant, but generally less reactive, organic compounds: carbon dioxide (370 parts per million, ppmv, 10−6 or μmol/mol), methane (1.8 ppmv) and carbon monoxide (0.15 ppmv), which have been discussed in detail elsewhere. Unfortunately various terms have been used in the literature to describe the subset of diverse carbon-containing gases under circa 10 ppbv. One of the first was nonmethane hydrocarbons (NMHC), which was employed originally to distinguish alkanes, such as ethane, propane and butane, from methane. However, strictly speaking the word ‘hydrocarbon’ indicates a molecule containing only carbon and hydrogen atoms, and therefore this term appears to exclude oxygenated species such as alcohols, carbonyls and acids as well as organic compounds containing other heteroatoms, such as nitrogen or sulphur. In an attempt to embrace all the species relevant to atmospheric chemistry, a further term ‘volatile organic compound’ (VOC) was coined, although there is no general quantitative definition of what VOCs are. The Environmental Protection Agency (EPA) in the United States has defined VOC as any compound that participates in atmospheric photochemical reactions; however, there have been subsequent attempts to give a more quantified definition. The result is that VOCs are considered to be those organic compounds having a vapour pressure greater than 10 Pa at 25◦ C, a boiling point of up to 260◦ C at atmospheric pressure, and 15 or less carbon atoms. The remaining compounds are designated as semivolatile organic compounds (SVOCs). This segregation emphasises the volatile gas phase species from those that partition to the aerosol phase which is reasonable since the later undergo different transport and chemistry (see Sections 1.3 and 1.4). Recently, however, the definition has become blurred by the use of OVOCs to specifically identify the oxygenated VOCs. For this chapter we prefer not to draw a divide through the continuum of

Volatile Organic Compounds in the Atmosphere

Stratospheric chemistry

Upper tropospheric chemistry

NO3 HO O3

Aerosol

N

VOC

Oxidation

siti

on

Advection

Remote region chemistry

Wa sh ou t Wet and dry deposition

olo

Bi

(Br, Cl)

Photochemical products Oxidation

uc

De

po

n

tio lea

C o nv e c ti o n

2

al

gic ke

ta

up

CO2 + H2O

Figure 1.1 Sketch of the various processes which determine the fate of VOC in the atmosphere. The individual processes are discussed in the text and the individual classes of VOC in the following chapters of this book.

compounds and instead use the term ‘trace organic compounds’, referring to the dictionary definitions of organic (designating carbon compounds) and trace (an extremely small amount). Despite being found at extremely low concentrations, trace organic compounds have profound effects in the atmosphere (see Figure 1.1). On the one hand, they are the ‘fuel’ which keeps oxidative atmospheric photochemistry running. Therefore, their sources, sinks and atmospheric residence times are the subject of much current research (see Sections 1.2–1.4). To investigate organic trace gases in the atmosphere it is essential that accurate concentration measurements and careful modelling studies are made (see Section 1.5). In addition to influencing local, regional and even global photochemistry, several such compounds have a potential impact on climate, both due to their properties as greenhouse gases and due to their ability to form aerosol particles on oxidation (see Section 1.6). Organic trace gases can be used as tracer compounds to investigate reactions which are not directly accessible to current measurement techniques or as probes to ‘visualise’ transport processes in the atmosphere or across atmospheric boundaries. Many open questions remain in this field of research, and some of the future challenges in the field are summarised in Section 1.7. The intention here is to provide an up-to-date, referenced overview of the field emphasising the recent progress made in an exciting and rapidly developing area of research. Recent and review-type references have been preferentially cited along with key older articles so that the interested reader may quickly access more detailed information. The authors would

Volatile Organic Compounds in the Atmosphere: An Overview

3

like to point out that the articles cited here represent a tiny fraction of a vast and widespread literature database. We hope that any omissions of particular works by colleagues will be forgiven in the interests of brevity.

1.2

Sources

Almost everything we do in daily life results in the release of organic species to the atmosphere. Driving a car (Fraser et al. 1998), painting the house (Fortmann et al. 1998), cooking (McDonald et al. 2003), making a fire (Andreae and Merlet 2001), cutting the grass (Fall et al. 1999; Kirstine et al. 1998) and even breathing (Barker et al. 2006; Phillips et al. 1999) – all of these processes result in the emission of organic compounds such as carbonyls, alcohols, alkanes, alkenes, esters, aromatics, ethers and amides. In addition to emissions from human activities, the Earth’s vegetation naturally releases huge amounts of organic gases into the air. As plants assimilate carbon dioxide into biomass through photosynthesis, a fraction of this carbon leaks out in to the atmosphere, predominantly in highly reduced forms such as isoprene and terpenes (Fuentes et al. 2000; Guenther 2002; Kesselmeier et al. 2002). Exactly which compounds are emitted from a particular plant, and how much of each, depends on the age and health of the vegetation as well as the ambient temperature, moisture and light levels (Guenther et al. 1995; Kesselmeier and Staudt 1999). Both plants and invertebrates have been shown to use emission of specific organic species into the air for signalling (Greene and Gordon 2003; Krieger and Breer 1999). Examples of elaborate chemical mimicry have been found in have been found in insects (e.g. Cremer et al. 2002) and amongst plants to deter attack by herbivores (Kaori et al. 2002; Kessler and Baldwin 2001). While the natural world uses air as a communication medium, man often uses it as a repository for waste products. The emission rates and their associated uncertainties for VOCs from several source categories are summarised in Table 1.1. The anthropogenic contribution to organic emissions in the atmosphere is dominated by the exploitation of fossil fuels (coal, oil and gas). Approximately 100 TgC/year was estimated to be emitted from ‘technological’ sources and 150 TgC/year from all anthropogenic sources including biomass burning (Müller 1992). Coal production mainly leads to methane emission, but minor emissions of ethane and propane are also present. Liquid fossil fuel production, storage and distribution result in a larger variety of organic gas emissions to the atmosphere. Crude oil production platforms are strong point sources of hydrocarbons such as methane, ethane, propane, butanes, pentanes, hexanes, heptanes, octanes and cycloparaffins (McInnes 1996). The major sources from processing liquid fossil fuels are catalytic cracking (0.25–0.63 kg/m3 of feed), coking (about 0.4 kg/m3 of feed) and asphalt blowing (about 27 kg of VOC/m3 of asphalt) (Friedrich and Obermeier 1999). Furthermore, so-called fugitive emissions can occur from leaks and evaporation from all types of equipment and installations. Evaporative emissions are estimated to be 2.9 kg/t of fuel at service stations (McInnes 1996) and are familiar to anyone who has filled a car with gasoline/petrol. Petrochemical products typically contain a limited number of compound classes (e.g. acyclic alkanes, cyclic alkanes, monoaromatics, diaromatics) each consisting of a very large number (tens of thousands) of individual homologues and isomers (Schoenmakers et al. 2000). Major products of the complete combustion of fossil fuels are carbon dioxide and water. However, in practice, combustion leads to CO and organic gas by-products, mainly due to

4

Volatile Organic Compounds in the Atmosphere

Table 1.1 Overview of important sources and global annual emission rates of selected groups of VOC per year Emission rate

Uncertainty range

Fossil fuel use Alkanes Alkenes Aromatic compounds

28 12 20

15–60 5–25 10–30

Biomass burning Alkanes Alkenes Aromatic compounds

15 20 5

7–30 10–30 2–10

Terrestrial plants Isoprene Sum of monoterpenes Sum of other VOC

460 140 580

Oceans Alkanes Alkenes Sum of anthropogenic and oceanic emissions Alkanes Alkenes Aromatic compounds

1 6

200–1 800 50–400 150–2 400 0–2 3–12

44 38 25

Terrestrial plants

1 180

Total

1 287

lack of oxygen, imperfect air/fuel mixing and inappropriate combustion temperatures. Tailpipe emissions from gasoline passenger cars with and without three-way catalyst are estimated, respectively, as 0.68 and 18.92 g of HC/kg of fuel, whereas passenger diesel cars (produced after 1996) emit about 1.32 g of HC/kg of fuel. Similarly, diesel heavyduty vehicle emissions are estimated as 5.4 g of HC/kg of fuel. Exhaust gas emissions from motor vehicles strongly depend on parameters such as vehicle speed, motor load and engine temperature. The predominate emissions for gasoline and diesel combustion engines are −C5 for gasoline cars and methane for diesel engines), C2 − −C5 olefins, ethyne, paraffins (C1 − aromatic hydrocarbons (BTEX and C9 aromatics), aldehydes (formaldehyde, acetaldehyde, acrolein, benzaldehyde, tolualdehyde), ketones (acetone) and others (mainly high molecular weight paraffins). A smaller emission contribution of anthropogenic gases comes from the solvents industry, and global inventories of these anthropogenic emissions have been compiled (Friedrich and Obermeier 1999; Olivier et al. 1999). These ‘anthropogenic’ emissions are discussed in detail in Chapter 2 of this book. A further strong source of global emissions is from burning of biomass, and giant smoke plumes can nowadays be seen easily on satellite images, especially in the tropics during the dry season (September–October). These emissions are the most difficult to assess as sources, as they are highly dependent on fuel type, humidity and burn rate amongst other

Volatile Organic Compounds in the Atmosphere: An Overview

5

factors (Lobert et al. 1990). Spatial and temporal variability further complicates global budget assessments, and satellite measurements are now being used to monitor the size and location of burning regions (e.g. Duncan et al. 2003). Most burning occurs during human-initiated land clearance but a large component also comes from the domestic use of biomass fuels (Levine 2003). A comprehensive summary of organic gas emissions from biomass burning relative to CO2 has been made recently (Andreae and Merlet 2001). On a global scale, the total amount of reactive biogenic emissions is not well established, although recent estimates indicate that c. 1300 TgC/year are emitted (Guenther 2002). The strongest biogenic emission is thought to be isoprene (C5 H8 ), followed by the less specified so-called other reactive biogenic compounds that are mainly oxygenated compounds and monoterpenes. Biogenic sources in total are considered to be approximately ten times larger than the sum of anthropogenic emissions including fossil fuel emissions and biomass burning (Guenther 2002; Muller 1992; Olivier et al. 1999). In comparison to terrestrial sources, emissions from the ocean are less well constrained although several important species are known to have a predominately marine source dimethyl sulphide (DMS, see Section 1.6) (Groene 1995) and methyl iodide (Lovelock 1975). A relatively small amount of organics is emitted from the ocean in the form of alkanes and alkenes, c. 5 TgC/year (e.g. Broadgate et al. 1997; Ratte et al. 1993). Recently global oceanic isoprene emissions have been estimated from satellite-derived chlorophyll map and laboratory studies 0.1 TgC/year (Palmer and Shaw 2005), that is, much smaller than the terrestrial source (c. 500 TgC/year). However, the surface ocean has been shown recently to be a massive reservoir for oxygenated organic species (Singh et al. 2003; Williams et al. 2004). Furthermore, a recent study of aerosols at a coastal site in Ireland (O’Dowd et al. 2004) showed that the organic fraction contributes significantly (63%) to the sub-micrometer particle mass of aerosols collected over the North Atlantic Ocean during phytoplankton bloom periods. Biogenic emissions in general are discussed in Chapter 3 of this book, while Chapter 4 includes a section concerning the biogenic formation of OVOCs. From a global perspective, geographical location and season determine the relative importance of anthropogenic and biogenic emissions: biogenics are emitted mostly in the tropics whereas most anthropogenic emissions occur in the northern hemisphere between 40◦ N and 50◦ N. All these diverse organic emissions are broken down in the atmosphere into a wider array of partially oxidised species (Atkinson 1994; Atkinson and Arey 2003; Jenkin et al. 1997) and many thousands of gases have been detected in the atmosphere, from the tropics to Antarctica (Ciccioli et al. 1996; Zimmerman et al. 1988).

1.3

Sinks

Since the concentrations of organic trace species do not all simply increase with time there must logically be one or more removal processes (here termed sinks) acting on these compounds. The most important sink for organic trace gases in the atmosphere is chemical oxidation in the gas phase by the hydroxyl radical HO (or to a lesser extent O3 , NO3 and halogen radicals) (Atkinson 1994; Jenkin et al. 2003; Saunders et al. 2003). Certain gas phase organic compounds in the air can absorb sunlight and thereby photolyse to smaller fragments. Some compounds can be efficiently removed physically by dry deposition to surfaces such as vegetation (Doskey et al. 2004; Muller 1992) or aerosol (Cousins and

6

Volatile Organic Compounds in the Atmosphere

Mackay 2001); or removed by wet deposition in rain (Fornaro and Gutz 2003; Kieber et al. 2002). The gas phase oxidation of organic compounds in air is mostly initiated by the HO radical, with carbon dioxide and water being the final products. In this way atmospheric oxidation is analogous to combustion. Using an everyday example as an analogy, when a cigarette lighter is lit, the hydrocarbon butane burns directly in the flame to form H2 O and CO2 . When the flame is not ignited, then the escaping butane gas is oxidised in the air to the same products, only much more slowly and via many other intermediates. The intermediate oxidation products may have lower vapour pressures, higher polarity or absorb light better than the precursors, making the intermediate products potentially more susceptible to physical removal or photolysis. An alkane must be larger than C20 to be adsorbed onto solid particles (Bidleman 1988), but much smaller multi-functional organic compounds, such as oxalic acid, more readily adsorb and are commonly found on aerosols (Mochida et al. 2003). Further oxidative transformation of these species on the aerosol is also possible (Claeys et al. 2004a; Noziere and Riemer 2003). The overall rate of removal of an organic species from the atmosphere can be derived by summing the reaction rates with radical species, rates of photolysis and the wet and dry deposition rates. From this we may determine the atmospheric lifetime of a species (see spatial distribution section). The rate of reaction of HO with many individual organic compounds under terrestrial conditions is well established from laboratory experiments as a function of temperature and pressure (http://www.iupac-kinetic.ch.cam.ac.uk/ and Mannschreck et al. 2002). Table 1.2 shows the atmospheric lifetime of several commonly measured VOCs with respect to OH, with lifetimes varying from months to minutes. Likewise, global photolysis rates can be calculated for many compounds from laboratory absorption cross section and quantum yield measurements (http://www. iupackinetic.ch.cam.ac.uk/). These rates can be profoundly influenced by clouds and this in turn can affect trace gas concentrations (Tie et al. 2003). The wet and dry deposition rates for organic compounds are highly variable and are generally empirically determined in the field. Generally, organic compounds measured at high and invariable concentrations in the atmosphere are less efficiently removed (Junge 1974). Relationships between the variability of organic gas measurements and their rate of removal by HO have been derived (Jobson et al. 1999; Williams et al. 2000) and exploited to derive HO trends. If a long-lived and hence well-distributed organic compound is known to react predominately with HO, and its emission and HO reaction rate are known, then the global HO concentration can be theoretically estimated. Initial attempts based on methyl chloroform indicated large changes in HO concentrations from 1978 to 2003 (Prinn et al. 2001, 2005) with a maximum in 1989 and a minimum in 1998, although other recent evidence suggests that uncertainty in the temporal and spatial emission pattern of methyl chloroform complicates such trend analysis (Krol et al. 2003). Direct biological uptake can also be an effective atmospheric removal process for some organic species (Kesselmeier 2001; Kuhn et al. 2002). The rate of uptake is dependent on the ambient concentration, being strongest when ambient concentrations are high. Compensation points have been deduced for plants, which mark the crossover point between emission and uptake. A surprising recent discovery is that peroxy acetyl nitrate (PAN), an anthropogenic secondary oxidant like ozone, can also be taken up by plants (Sparks et al. 2003). This is an important development for the atmospheric nitrogen cycle as well as the organic species PAN.

Volatile Organic Compounds in the Atmosphere: An Overview

7

Table 1.2 Overview of average tropospheric lifetimes of VOC compound groups and some selected VOCs as examples. Lifetimes are given for an average OH concentration of 6 × 105 per cm3 and an average ozone concentration of 7 × 1011 per cm3 (about 30 ppb)

1.4

Compound

Average lifetime

Alkanes Ethane Propane n-Pentane

Months–days 2.5 months 2.5 weeks 4 days

Alkenes Ethene Propene 1-Butene

Days–hours 1.5 days 11 h 10 h

Cyclic compounds Cyclopentane Methylcyclohexane Cyclohexane

Days–hours 4 days 2 days 3h

Aromatic compounds Benzene Toluene 1,3,5-Trimethylbenzene

Weeks–hours 2 weeks 2 days 7.5 h

Biogenic compounds Isoprene α-Pinene Limonene

Hours–minutes 3h 4h 30 min

Atmospheric distribution

Following emission, volatile organic species are distributed by atmospheric transport processes while undergoing the removal processes described in Section 1.3. The relationship between the atmospheric removal rate (lifetime) of the compound and the average mixing times for the various sections of the atmosphere together determine the extent to which a compound is globally distributed. Atmospheric mixing is physically impeded across the boundary layer temperature inversion (0.5–2 km), the tropopause (10–15 km) and the intertropical conversion zone (ITCZ, 10◦ S–10◦ N), and as a result strong gradients in organic species can develop across these atmospheric interfaces. Typical exchange times are 1–2 days for air to mix vertically out of the boundary layer, 2 weeks to a month for air to be advected zonally around the northern or southern hemisphere, about 1 year for interhemispheric exchange and 4–6 years for an exchange between troposphere and stratosphere. Chemical lifetimes, defined as the time for a chemical concentration to decay to 1/e of its initial value, varies from minutes to hours (terpenes and isoprene), through days to weeks (acetone, methanol, propane), years to decades (methyl chloroform, HCFC 134a), and up to hundreds

8

Volatile Organic Compounds in the Atmosphere

of years for chlorofluorocarbons (CFC 11 and CFC 12) whose lifetime is determined by the mixing rate into the stratosphere. Short-lived compounds, such as the biogenic species isoprene (CH2 C(CH3 )CHCH2 ), show strong atmospheric gradients within the boundary layer (0–2 km), whereas longer-lived compounds, such as CFC 113 (lifetime c. 8 years), are better mixed and only show strong gradients between the hemispheres (Boissard et al. 1996; Bonsang and Boissard 1999; Rudolph 1998). Some compounds are more or less uniformly distributed in the troposphere (CFC 12, lifetime 79 years, no remaining sources), only showing concentration gradients in the stratosphere. A large archive of aircraft measurements taken at various locations over the globe is available at http://www-gte.larc.nasa.gov/. There are numerous examples in the literature of regional scale advection where organic pollutants found in remote locations have been linked to distant pollution sources by use of back trajectories (e.g. Blake et al. 1996; Traub et al. 2003). Intercontinental pollution events have been reported (Price et al. 2004; Stohl and Trickl 1999) and trajectories have even been used to track southern hemispheric biomass burning, through the ITCZ to the upper troposphere of the northern hemisphere (Andreae et al. 2001). Secondary photooxidants, such as ozone and PAN, that form en route have also similarly been identified in plumes emerging from urban centres (e.g. Rappengluck et al. 2003). Interestingly, there is growing evidence to suggest that migrating birds use chemical gradients as an olfactory aid to navigation (Wallraff 2001, 2003). Where the atmosphere is in contact with the Earth, organic species can interact with the various surfaces, such as snow, soil and water (Ballschmiter 1992). Within these media further production or removal mechanisms may exist such as bacterial uptake, enhanced photolysis (Dominé and Shepson 2002; Klán et al. 2003) or biological production. Such processes will affect the lifetime of these species and hence their global distribution. Some higher molecular weight organics with considerably lower vapour pressures tend to partition predominantly to aerosols following release. When such a compound is unreactive, as with persistent organic pollutants (POPs), which are emitted through incomplete combustion or pesticide use, the lifetime of the transporting aerosol will then determine the distribution of this compound. Examples of such compounds include polyaromatic hydrocarbons (PAHs, e.g. Mastral and Callén 2000), Polychlorinated biphenyls (PCBs) and polychlorinated dibenzo-p-dioxins (PCDDs). Whether in gas form or as particles, these compounds can be transported long distances from source regions (Patton et al. 1991). The distribution of the long-lived semi-volatiles is markedly different to that of the volatiles, and with time through repeated volatilisation and adsorption, such compounds tend to concentrate in polar regions (Burkow and Kallenborn 2000) in a manner that could be likened to a global distillation (from the tropics to the poles). Some of these compounds are toxic (Walker 2001a) and can bioaccumulate through the food web (Tanabe et al. 1984), posing a risk to human health and the environment (UNEP 2001). While the boundary layer (0–2 km) tends to be turbulent, the so-called free troposphere above is less well mixed. In addition to the slow process of diffusion, organic gases may be distributed in the atmosphere by meteorological events such as convection (Collins et al. 1999) and via lifting by frontal systems (Purvis et al. 2003). The overall distribution of the organic species varies with latitude and season as a function of the source and sink strengths, as well as prevailing meteorology (Bonsang and Boissard 1999; Singh and Zimmerman 1992). Certain photochemical products, such as organic nitrates (e.g. PAN or alkyl nitrates), have a hemispheric concentration that is maximum in the spring. This has

Volatile Organic Compounds in the Atmosphere: An Overview

9

been explained as the photochemical optimum between the high precursor source and low photochemical sink in wintertime, when PAN precursors are accumulated, and the high photochemical sink in summer (Penkett 1983). In the early years of atmospheric research, it was assumed that after several days to weeks the atmosphere would have effectively removed an organic pollutant, based on the atmospheric lifetime of alkanes. Recently, however, from measurements made between 1 and 13 km over the remote Pacific Ocean, far from source regions, it was shown that volume mixing ratios of oxygenated organic species are some five times higher than those of the NMHC, alkanes and alkenes (Singh et al. 2001, 2004). Similar high mixing ratios of oxygenates and compound diversity have been reported in other airborne studies (Crutzen et al. 2000), in urban centres (Lewis et al. 2000) and in continental outflow from Asia (Jacob et al. 2003; Lelieveld et al. 2001) and Europe (Salisbury et al. 2003). These results concur with earlier theoretical work on the oxidation of organic compounds (Calvert and Madronich 1987; Madronich et al. 1990). Our views about the distribution, sources and role of reactive organic species in the atmosphere are currently being revised rapidly.

1.5

Measurement tools

The human nose is particularly sensitive to several chemical groups (Cain 1979; Firestein 2001). Familiar examples include forests, which emit terpenes (e.g. Geron et al. 2000; Isidorov et al. 2000); oil refineries, which emit aromatic compounds and alkenes (e.g. Doskey et al. 1999); fish markets, which emit amines (Morita et al. 2003) and freshly cut onions, which emit sulphur compounds (e.g. Ferary and Auger 1996). While human subjects are widely used in odour identification studies (Ferreira et al. 2003; Walker 2001b), the nose’s response is inherently subjective (Molhave et al. 1991) and difficult to quantify. Therefore, to investigate the atmosphere quantitatively, researchers have employed a variety of sensitive and specific sensors, including mass spectrometers, flame-ionisation detectors, electroncapture detectors, optical absorption, chemiluminescence and atomic emission detectors (e.g. Apel et al. 1998; Helmig 1999; Kormann et al. 2002; Sigrist 2003). In many cases such detectors are coupled to pre-separation devices, for example, a gas chromatograph so that individual gases may be isolated prior to detection and a single specific compound can be measured. Both animal- and plant-type biological detectors have also been deployed for detection of certain molecules. In some studies the amputated sensory antennae of small insects have been connected into measurement devices (e.g. Murlis et al. 2000) and elsewhere the leaves of plants have been analysed for long-term exposure statistics (Hiatt 1999). Ideally for air studies a detectors should also be capable of measuring the huge range of concentrations in the atmosphere. High mixing ratios of several tens of ppbv (nmol/mol) can be found for alkanes and aromatics in polluted urban areas urban areas (Derwent et al. 2000), while halons must be reliably measured at only 0.045 pptv (pmol/mol) (Fraser et al. 1999). The recently reported compound SF5 CF3 was first detected at 0.005 pptv (Sturges et al. 2000). This means that if only 200 tonnes of such material would be emitted anywhere in the world it would, in time, be detectable by this instrument. Global networks of detectors are in currently place to routinely monitor changes in greenhouse gases (Prinn et al. 2000). Much of our atmospheric knowledge to date has been driven by what can be reliably measured and how fast (Roscoe and Clemitshaw 1997). Although the first atmospheric

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Volatile Organic Compounds in the Atmosphere

research on organic trace gases (specifically PAN) was made using infrared spectroscopy (Stephens 1961; Stephens et al. 1956), the following 30–40 years of research on atmospheric organic gases have been dominated by gas chromatography (e.g. Darley et al. 1963; Helmig 1999) coupled to some form of detector. Samples have been either introduced directly into the instrument in the field or collected in pressurised steel canisters, absorbent packed cartridges or filters for later analysis in the laboratory. The alkanes (major components of fossil fuel) were one of the first and most widely investigated subset of the reactive organic species (Blomberg et al. 2002). This is because these fully saturated compounds do not interact strongly with most inlet materials or collection vessel surfaces, and the long established technique of gas chromatographic separation with flame ionisation detection has allowed widely available quantitative analysis (Helmig 1999). Many oxidised gases are more difficult to quantify as they may stick to surfaces, thermally decompose or may even be produced in measurement systems (Bates et al. 2000; Helmig et al. 1996; Kelly and Holdren 1995; Lestremau et al. 2001; Li-Jones et al. 2001; Tanner 2003). These techniques are sensitive and specific, but due to the predetector sample separation, these are limited in sampling frequency. Recently, several important new advances have been made in analytical techniques. These have permitted more organic species to be measured, such as high precision and sensitivity so as to enable δ 13 C isotopic ratios to be determined in organics at mixing ratios below ppbv levels (Rudolph et al. 1997) (see Chapter 10 of this book). Higher frequency measurements have been made possible with chemical ionisation mass spectrometry (CIMS) through use of proton transfer reactions (Hewitt et al. 2003; Lindinger et al. 1998; Williams et al. 2001) (see Chapter 11 of this book), or by other chemical ionisation techniques (Heeb et al. 1999; Leibrock et al. 2003). Several of these high frequency methods have been further developed to measure emission fluxes directly (Bowling et al. 1998; Karl et al. 2004; Warneke et al. 2002). Furthermore, improvements in the field of gas chromatography (e.g. multi-dimensional gas chromatography or comprehensive chromatography, Phillips et al. 1985) have delivered considerable improvements in compound separation, identification and sensitivity (see Chapter 12 of this book). The physical separation of even enantiomeric monoterpenes, or optical isomers is also now possible from ambient air (Yassaa and Williams 2005). With the arrival of this new generation of measurement systems, more species and timescales are accessible and a new golden age of discovery for field measurement has begun. Researchers are now exploiting these latest techniques on planes, ships, balloons and ground sites to establish the global budgets of a wide range of organic species.

1.6

Modelling tools

A variety of numerical models is available today to simulate chemistry and transport in the atmosphere from the level of box models to three-dimensional chemistry and transport models. To simulate atmospheric chemistry in detail, models need to deal with VOCs which play a significant role in all reaction cycles in the atmosphere. In this regard the major input data they require are (a) the emission inventories describing the primary emission of VOC including their specific source compositions and their spatial and temporal variations, (b) the oxidation chemistry of VOC including the kinetics as a function of temperature and pressure and (c) for certain species a consideration of other significant loss processes such as

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dry and wet deposition. Understandably, due to the huge number of organic compounds in the troposphere, both a complete emission inventory of all possible compounds with the necessary resolution in time and space and a complete coverage of all possible chemical reactions including those of secondary reaction products will never be available. There are only a few comprehensive emission inventories available and these cover the most important compounds relevant for atmospheric chemistry. The global distribution and source strengths of anthropogenic NMHC are usually taken from the Emission Database for Global Atmospheric Research (EDGAR V2.0) database (Olivier et al. 1996, 1999). Detailed information on this database and access to the data is available on http://www.mnp.nl/edgar/. The EDGAR database details sources of fossil-fuel-related activities, biofuel combustion, industrial production and consumption processes (including solvent use) on a per country basis, land-use-related sources, including waste treatment, partially on a grid basis and partially on a per country basis; and natural sources on a grid basis. The database can be used to generate global, regional and national emission data in various formats. For all compounds the reference year is 1990, except for halocarbons, for which 1986 is the reference year. In 2001, version 3.2 was released which comprises emissions by region and source for the period 1990–5. However, organic compounds emissions given in the EDGAR database are known to have large uncertainties in both magnitude and distribution of the emissions. An inventory of biomass burning and natural VOC emissions can be found in the Global Emissions Inventory Activity (GEIA) database. Details of this database can be found on http://www.geiacenter.org/. The biogenic VOC (BVOC) dataset consists of three files that cover isoprene, terpenes and ‘other’ NVOC data. This database provides BVOC emission measurements and modelling parameters and, in addition, an enclosure database that summarises information from literature and identifies the plant species and the BVOC studied, including enclosure and analytical techniques and other parameters. Presently, about 1 800 plant species from which BVOC emissions have been studied are documented. As will be described in the following chapters, for individual VOC compound classes the chemistry of VOC can be quite complex. Two main approaches are used to simulate the complex chemistry of VOC using numerical models. One approach is to significantly reduce the number of organic compounds and hence the complexity of the corresponding reactions. Three methods are used to reduce the number of organic reactions, namely the carbon bond mechanism, the surrogate species method and the lumped species method. In the carbon bond mechanism, individual organic compounds are segregated into one or more bond groups that have a similar chemical reactivity (cf. Gery et al. 1988). With the surrogate species method all VOC of similar reactivity are grouped together. The rate coefficient of each of these compounds is then set equal to that of one particular compound (cf. Atkinson et al. 1982). In the lumped species method, VOC are grouped by their reactivity towards reactions with HO radicals. The rate coefficient is determined by taking the mole fraction weighted average of the reaction coefficient of each compound of the lumped group. Currently the most widely used lumped mechanism is the Regional Acid Deposition Model (RADM) (Stockwell et al. 1990). This mechanism contains 158 chemical reactions and 63 gaseous compounds. Besides primarily emitted inorganic compounds and 16 organic compounds or compound groups, respectively, RADM also includes photochemically produced compounds. This mechanism has meanwhile been updated into a version named

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Volatile Organic Compounds in the Atmosphere

Regional Atmospheric Chemistry Model (RACM) (Stockwell et al. 1997) and includes reaction schemes for isoprene and the group of monoterpenes. The RACM mechanism again has been improved (Regional Atmospheric Chemistry Model-Mainz Isoprene Mechanism, RACM-MIM) by implementing an explicit chemistry of isoprene (Geiger et al. 2003), the most important BVOC (see Chapter 3). Today RACM is widely used in a variety of mathematical models, ranging from box models to three-dimensional chemistry and transport models. However, the mechanism is also available as a stand-alone model for simulating tropospheric chemistry at typical ambient trace gas concentration levels. The aforementioned methods of condensing VOC reactions have, of course, some disadvantages. The results of a model investigation of a single hydrocarbon oxidation are likely to be erroneous. Investigation of the RACM simulation showed, for example, that for a simulation of the oxidation of some individual VOC, such as branched alkanes, the error can be considerable. Therefore, other approaches have been developed that try to describe the complex chemistry of organics as explicitly as possible, leading to a vast number of chemicals and reactions which have to be taken into account. For example, the widely used Master Chemical Mechanism (MCM) is a detailed chemical mechanism containing 12 600 reactions and 4 500 chemical species (Jenkin et al. 2003). This mechanism includes the complete tropospheric oxidation of 124 VOC. The VOC which are degraded in this mechanism were selected on the basis of available emission data and provide approximately 90% mass coverage of the emissions of uniquely identifiable chemical species. The majority of the degradation schemes have been constructed using the methodology described by Jenkin et al. (1997). A review and update of the ideas behind the mechanism as well as recent developments can be found on the corresponding web page (http://mcm.leeds.ac.uk/MCM/).

1.7

How organic species affect the atmosphere

In the 1950s Haagen Smit and co-workers showed that the oxidation of organic species in the presence of NOx and sunlight can form ozone (Haagen Smit 1952). Ozone, which is toxic to humans and plants, has become a major air quality problem in cities and larger areas such as the Mediterranean (Lelieveld et al. 2002) and the south-eastern United States (Solomon et al. 2000). Ozone control strategies adopted in the 1970s were initially unsuccessful due to an underestimation of natural organic emissions in the initial models (Trainer et al. 1987). However, more recent emission controls applied to cars (including NOx and organic species reductions) have reduced regionally produced ozone (Derwent et al. 2003). The capacity of the troposphere to oxidise emissions is also dependent on the amount of organic compounds present. Reaction with the main atmospheric oxidant, the HO radical, is the primary loss mechanism of organics from the atmosphere (see Section 1.3). While the initial reaction is a sink, subsequent oxidation steps may be a source of HOx (HO and HO2 ), making the global effect of organic species complicated. In cities where NOx concentrations are high, increasing concentrations of organics increase the ambient HO. However, in most of the free troposphere HO production is not limited by organics but rather by NOx , and increasing organic concentrations generally decreases ambient HO under these conditions (Wang et al. 1998). In the upper troposphere, where water concentrations are low (<100 ppmv), organic species may provide the main source of HOx radicals. For example, acetone is a source of HOx when photolysed in the dry upper troposphere

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(McKeen et al. 1997) and can lead to strong ozone formation. Similar effects are produced by organic peroxides and aldehydes (Jaeglé et al. 2000). At night the organics may also provide an important source of HO radicals (e.g. Platt et al. 2002) through the reactions of alkenes with ozone, and in oceanic regions, organohalogens can provide ozone-depleting halogen radicals such as Br (Platt and Honninger 2003), which can link with inorganic halogen cycles (Sander et al. 2003). The global impact of organics on ozone and the hydroxyl radical have been investigated in models (Houweling et al. 1998; Wang et al. 1998). Sensitivity studies in global models indicate that removing hydrocarbon emissions gives modest decreases in global ozone (<15%) and relatively small increases in global mean HO (<20%). These studies highlight the role of organic nitrates such as PAN in the distribution of NOx and hence HOx . These nitrates form where hydrocarbons are oxidised in the presence of NOx (Roberts 1990). The most abundant nitrate PAN is a lachrymator and largely responsible for the sore eyes experienced in smog. Having a longer lifetime than NOx , organic nitrates may be transported much further from the pollution sources before decomposing to release NOx again. In this way these compounds function as a long-distance transport mechanism for NOx and in so doing they influence the global oxidation budget (Singh and Salas 1989). The effect of organic species on the global distribution of ozone and HO is a key area of atmospheric research. In the troposphere photooxidation of organic gases in the presence of high NOx (NO and NO2 ) concentrations acts to produce ozone; however in contrast, other trace organic gases can act to destroy ozone in the stratosphere. Prior to man’s proliferation on the planet, long-lived naturally produced organohalogens, such as methyl chloride, represented the most important mechanism for chlorine transport to the stratosphere and hence ozone destruction (Anderson 1990). More recently man-made chlorinated, brominated and fluorinated hydrocarbons have been shown to deliver significant additional quantities of chlorine and bromine into the stratosphere (Rowland 1990) and be the cause of the ‘ozone holes’ observed over the polar regions (Molina 1988). As a result of legislation, the overall tropospheric abundance of halogen from halocarbons is now decreasing (Montzka et al. 1996). However, emissions of certain species (e.g. Halons) persist because of a lack of suitable substitutes for critical uses such as fire extinguishants (Butler et al. 1998). In addition to the major organic greenhouse gases CO2 and CH4 , certain other organic species have been implicated in long term or climate effects. Perhaps the best-known example is DMS, a compound which is naturally emitted from the oceans, and that has been proposed as a potential negative feedback to climate warming (Charlson et al. 1987). The hypothesis is that a long-term warming of the oceans would produce more DMS emission, which following oxidation to SO2 and then sulphate would lead to more clouds and hence more reflection of incoming sunlight. In polluted regions, where organic species can provide large numbers of condensation nuclei for clouds, then a further radiative consequence of organics emerges. The larger number of nuclei means that the available water in the cloud is more widely distributed, causing the average droplet size to be smaller and the cloud as a result to be more reflective (Feingold et al. 2003; Platnick and Twomey 1994). More recently certain oceanic alkyl halides have been linked to cloud formation and hence to possible climatic effects (O’Dowd 2002). While all organic species are infrared active to some extent, their influence on the global radiative forcing depends on their absorption spectrum and atmospheric abundance (Harries et al. 2001). The CFCs and HCFCs are important in this

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Volatile Organic Compounds in the Atmosphere

regard. Therefore, organic gases, or the particulate products thereof, can cause a direct effect on climate forcing by reflecting or absorbing incoming light, or an indirect effect through modification of cloud albedo and lifetime. By influencing global ozone, organic species may through chemistry also affect the radiation indirectly (Ramanathan 1998; Wigley et al. 2002). Due to the wide range of physical and chemical properties of organic compounds, it is extremely difficult to assess their overall climate feedback effects, especially when the emissions are likely to be also changing as a function of time (Sanderson et al. 2003). When considering the climate effects we may also consider possible roles of organic species on our past climate. Atmospheric organic species have also been implicated in the beginning of life, which has influenced our climate profoundly. Small micelles and reverse micelles resulting from high molecular weight organics and water coating on aerosol have been proposed as photoreactors within which complex proteins may build up (Tuck 2002; Vaida 2002). On a larger scale, primordial ocean oil slicks of organic species have also been proposed as a means of climate change and important in the formation of the first proteins (Nilson 2002). It has also been speculated that other organic species (nitriles) provided a source of nitrogen to photosynthesising marine organisms in the ancient oxygen-free atmosphere of the Earth (Bange and Williams 2000). The trace organic species also have enormous potential for revealing the atmospheric chemical history of the Earth. The analysis of air trapped in ice cores for CH4 and N2 O has provided atmospheric information of the past 120 000 year. Individual organic species can be more specifically attributed to sources than long-lived species such as CH4 and N2 O. Early work on small cores dating back only c. 40 year have demonstrated this potential (Lee et al. 2003). That gas phase organic compounds and atmospheric aerosols are strongly linked has been established for some time (Went et al. 1967), and organic atmospheric aerosols have recently been reviewed in detail by Seinfeld and Pankow (2003) and Jacobson et al. (2000). The blue hazes and reduced visibility over forests are the result of enhanced scattering of blue light by particles similar in size to the wavelength of light (Watson 2002). These particles can be produced from gaseous organic precursors, which condense from the gas phase to form aerosols (Kavouras et al. 1998). In this respect, oxygenated species with low vapour pressures are favoured; examples include the photochemical products of biogenic emissions such as isoprene (2-methyltetrols), terpenes (pinic and norpinic acid) and sesquiterpenes (Bonn and Moortgat 2003; Claeys et al. 2004b; O’Dowd et al. 2002). Such nuclei may grow in size by coagulation with other particles and later through the condensation of other organic species onto the surfaces (Griffin et al. 1999; Kavouras and Stephanou 2002; Kulmala 2003). Recent evidence has shown that organic compounds absorbed into particles may undergo acidcatalysed reactions: oxidation, hydration, hemiacetal and acetal formation; polymerisation and aldol condensation (Claeys et al. 2004b; Iinuma et al. 2004; Jang et al. 2002; Kalberer et al. 2004; Tolocka et al. 2004) and chemical oxidation of organics such as isoprene can occur in the aerosol liquid phase through acid catalysed reactions with hydrogen peroxide (Claeys et al. 2004a). As the particle grows, the hydrophilic and light scattering properties of the particle can be affected by condensing organic gases or oxidation processes. In pristine conditions over the Amazon, it has been shown that hydrophilic organic species are a large fraction of wet season aerosol mass and they are predicted to significantly contribute to particle growth into cloud condensation nuclei (Roberts et al. 2002; Yu 2002). An organic layer has also been recently reported on marine aerosols (Tervahattu et al. 2002). In more

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polluted conditions, it has been shown that uptake of organic gases onto soot particles can change particle reflectivity (Saathoff et al. 2003), making light absorbing soot particles more reflective. The opposite can also be speculated, that organic species can make reflective ammonium sulphate particles darker and more light absorbing. The organics, therefore, play a critical role in determining the aerosol albedo. As precursors of cloud condensation nuclei, organic gases can be important in the formation of clouds. Pure water requires extremely high relative humidities (400% or more) in order to produce clusters that provide nuclei for further condensation (Pruppacher and Klett 1997). However, in the presence of aerosols, condensation can occur at conditions more readily found in the atmosphere. Sulphate is a very effective cloud condensation nucleus and hence much research has been focused on DMS and its oxidation products (Pandis et al. 1994). A wide range of anthropogenic and biogenic compounds have been tested for their effect on ice nucleating properties (e.g. Szyrmer and Zawadzki 1997). Amino acids have been shown to be particularly effective in this regard (Milne and Zika 1993), and while long-chain alcohols in monolayers can also promote ice formation (Gavish et al. 1990), some organic acids appear to delay activation (Shantz et al. 2003). There is also some evidence that organic gases can influence the shape of ice crystals (Hallett and Mason 1958), a parameter that is predicted to have a large effect on snowfall rate (Lohmann et al. 2003). In cities, organic aerosols and soot are also emitted directly, particularly from diesel exhausts. POPs such as PAHs (carcinogenic) and PCBs (toxics derived from burning plastics) are often associated with urban particles. Particles smaller than 10 μm (PM 10) can be effectively inhaled by humans, and correlations have been shown with mortality rates (Brunekreef et al. 1995). In the clean marine environment it has been shown recently that aerosols can be efficiently formed from iodine-containing organics, such as diiodomethane (Jimenez et al. 2003). Furthermore, following formation marine aerosol appears to be coated with organic fatty acids. This coating of organic surfactants could have important effects on the physical and chemical properties of the aerosol (Gill et al. 1983). The impact of ship emissions of organics on the atmospheric aerosol is also currently under investigation (von Glasow et al. 2003; Noone et al. 2000). A more detailed discussion of organic aerosols can be found in Chapter 9 of this book.

1.8

Open questions and future directions

The previous sections have outlined how huge emissions of organic gases enter the atmosphere and how oxidation, predominantly initiated by the HO radical, breaks down these gases to CO2 and water. Determining how much reactive carbon is in ambient air and comparing it to the individually measured compounds is an important ongoing task. A similar budgeting process has been successfully performed for atmospheric nitrogen species (Fahey et al. 1986). For the carbon compounds this job is made difficult by the presence of CO, CH4 and CO2 , whose concentrations dwarf those of the more reactive species. Some initial attempts have been made to measure the total carbon, termed Cy by chromatographic separation and then conversion of all reactive species to CH4 (Maris et al. 2003; Roberts et al. 1998). Comparisons to individually measured hydrocarbons are reasonable suggesting that most compounds making up the budget are being measured individually. However, the techniques adopted involve significant sample handling, where losses of some

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Volatile Organic Compounds in the Atmosphere

organic species could occur (Chung et al. 2003). Laboratory experiments and some as yet unpublished field studies have shown the potential of HO measurements to determine total reactivity (Jeanneret et al. 2001; Kirchner et al. 2001), but the size and complexity of the instrumentation used has until now precluded widespread use. Recently ground-based measurements made over a forest have shown that a considerable fraction of the measured reactivity could not be accounted for by speciated measurements and that the missing reactivity had a terpene-like emission profile (Di Carlo et al. 2004). Further forest-based ozone fluxes and possible unidentified photoproducts support this notion (Goldstein et al. 2004; Holzinger et al. 2005). Generally, global models have considerable difficulty simulating the chemistry over large forest areas (e.g. the Amazon rainforest) since the high fluxes from known reactive organic compounds, such as isoprene and monoterpenes, tend to consume the available HO radicals in the model. Clearly much more work must be performed to determine what fraction of the ambient organic trace gases are being measured by current techniques. If the total reactivity can be reliably determined, it can simplify photochemical modelling of organic species and aid the development of ozone control strategies. A further open issue is chemical oxidation in the polluted atmosphere. How well we understand the oxidation process can be generally assessed from the agreement between values from a theoretical model and direct measurement. Formaldehyde is considered a good parameter to compare since almost all organic species produce formaldehyde at some stage during gas phase oxidation. While measured and modelled comparisons of formaldehyde show good agreement in clean environments, comparisons in polluted air have consistently shown an underestimate in model values (Fried et al. 2003), the difference being up to a factor of 3–4 in the upper troposphere (Kormann et al. 2003). Assuming this is not a result of transport or source strength errors in the model, this suggests that the organic chemistry in such circumstances is more complex than currently thought. The ozone-forming potential of the organic species is therefore not well understood. Needed are comprehensive field measurement datasets of organic compounds including precursor and oxygenated products. These must be compared with explicit chemistry models (e.g. Calvert et al. 2002; Jenkin et al. 1997) to determine for which species real air oxidation chemistry differs from the theoretical oxidation pathways. A complicating factor in this research will be the determination of the gas-to-particle partitioning of the organic species and representation of multi-phase chemistry in models. The next generation of atmospheric chemical transport models will require realistic but tractable mechanisms for organic oxidation, and particulate formation and growth in order to calculate the radiative forcing of climate (Chung and Seinfeld 2002; Griffin et al. 2002). The role of the ocean in the budgets of organic species is surprisingly poorly understood. Many organic species are reported to be emitted from the ocean (including sulphurcontaining gases (Kettle et al. 2001), organohalogens (Carpenter et al. 2003), alkyl nitrates (Chuck et al. 2002)) while many other species are taken up (e.g. acetone, Warneke and de Gouw 2001; methanol, Galbally and Kirstine 2002). In reality, the ocean surface may be a highly variable source or sink for many compounds depending on the latitude, temperature, wind speed and biological composition of the surface water. Ocean emissions may also be dependent on aeolian input of trace elements such as iron and phosphorous. It has been speculated that for some compounds the ocean surface layer represents a giant reservoir of organic species exceeding the amount in the troposphere (e.g. for methanol; Galbally and Kirstine 2002; Singh et al. 2003; Williams et al. 2004). The ocean is able

Volatile Organic Compounds in the Atmosphere: An Overview

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to assimilate CO2 at rates per unit biomass five times greater than the largest terrestrial ecosystems, namely, the rainforests, but oceanic organic trace gas production is not well understood. Considering its size and potential importance, the ocean is surprisingly poorly characterised in terms of organic gases, although new measurement programme have recently been initiated (Duce and Liss 2002). Of particular importance to atmospheric chemistry is the characterisation of tropical waters, as emission or uptake in these regions can affect air that is subsequently rapidly convected to the upper atmosphere in the ITCZ. As new instrumentation is deployed in new locations, new compounds are discovered in the air. Important sources and sinks of organic species are continuously being uncovered (e.g. toluene from plants, Heiden et al. 1999; methyl chloride from plants, Yokouchi et al. 2002). Compilation of global emission inventories is then by definition an ongoing task. Production of an emissions inventory is a long and laborious process, and the ‘current’ version is, therefore, out of date as soon as it is published, leading to large discrepancies between model and measurement even for anthropogenically emitted species (Gros et al. 2003). From the last section it is clear that an understanding of the sources, sinks and chemistry of organic species in the atmosphere is important in predicting future global change. If the Earth warms, as it is predicted to do, then we may expect the distribution of organics to be affected through concomitant increases in temperature, convection and changing vegetation patterns. How biogenic emissions change as a function of the forecasted increases in CO2 will be particularly important, and early indications are that biogenic emissions increase in elevated CO2 conditions (monoterpenes, Constable et al. 1999; methane, Inubushi et al. 2003). However, ozone is also expected to increase, which may also have an effect on these emissions (Mclaughlin and Downing 1995). Predictions based on the Intergovernmental Panel on Climate Change (IPCC) recommended emission scenarios for organic and other species show increases in tropospheric ozone (Zeng and Pyle 2003) and changing radiation budgets (Wigley et al. 2002), which will lead to global air circulation change. These effects may in turn lead to changes in the hydrological and biogeochemical cycles. To offset warming effects of CO2 , a number of proactive methods have been suggested to reduce directly the atmospheric CO2 burden that are relevant to organic species. Methods such as intensive tree planting and fertilising the ocean with iron should be carefully vetted for side effects associated with organic emissions before implementation. For example, plantations of fast-growing trees in polluted areas such as black larch as carbon sinks, can lead to extreme local terpene emissions with important consequences for regional pollution. In a sense this can convert a global problem into a regional one. A similar unexpected outcome can occur in the Amazon, where oil palms are often planted following tree clearance. Since the oil palm is a strong emitter of biogenic reactive compounds, it is conceivable that man’s interference can increase rather than decrease emission rates from a region. Similar effects have been noted in the United States, for example, through proliferation of sweetgum in pine plantations, and it has been suggested that VOC emissions are increasing at 6% per decade in the United States (Purves et al. 2004). Experiments aimed at fertilising the ocean with iron in order to stimulate biological growth and hence CO2 uptake should be also carefully assessed for enhanced organic halogen- or sulphurcontaining emissions which can effect ozone chemistry and aerosol formation (Fuhrman and Capone 1991). In the future, the global population is expected to grow, and megacities will house greater proportions of the increasing population (Cohen 2003). Therefore, it will be increasingly

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Volatile Organic Compounds in the Atmosphere

important to control the local air quality in order to maintain human health (Beer 2001; Rodwin and Gusmano 2002). This will have to be achieved in the face of a long-term global upward trend in background ozone (Lelieveld et al. 2004; Prather et al. 2003). Considerable effort is being made to reduce the organic emissions of automobile engines through use of alternative fuels such as compressed natural gas (CNG) and liquid petroleum gas (LPG) as well as new engine technologies (Pearson 2001). Use of hydrogen fuel cell technology has been predicted to reduce CO and NOx emissions by up to 50% but its impact on climate forcing is dependent of the technology used (Schultz et al. 2003). The use of filters in combination with diesel fuel has been predicted to significantly improve air quality in the short term (Jacobson et al. 2000). Indoor pollution in cities is also likely to grow in importance. Increased use of terpene-based propellants and increasing background ozone, both combined with the construction of more airtight energy efficient buildings all lead to greater exposure to organic chemicals (Baker 1994; Carslaw 2003). Concerns are already being raised by some nations that other nations situated upwind are responsible for their deteriorating air quality. This echoes the smaller scale situation in the 1980s when cities on the US east coast complained that they could not comply with national air quality standards due to pollution built up in the west. Similarly, in the 1970s acid rain in Scandinavia was attributed to sulphur emissions in the United Kingdom. The issue of international pollution export (UNECE 1991) will almost certainly gain importance in coming years. Legal action is conceivable, where one country demands recompense from another for perceived health effects or tourism decline. Cost estimates for damage attributable to organic compounds have been made already: $1 100 per Mg-VOC (Rabl and Eyre 1998). The monitoring of exported emissions from other countries may also be used as an indirect way of spying on the upwind country. In theory, in the future it could be possible to determine whether published economic figures are realistic by comparing trends with emissions, or to determine whether certain processes are in operation. Emission trading has already been introduced for CO2 , despite the fact that the Kyoto protocol has only recently been ratified. Emission trading in other species is under discussion (Solomon 1999). In theory, limits could be imposed to trace organic species but this would require a much better understanding of emissions and would logically entail policing, which considering the short lifetimes and small concentrations of many compounds seems unlikely in the short term. In the near future investigation of such compounds will remain a task for the atmospheric research community. Data from a variety of new satellite platforms (Borrell et al. 2003) is unlikely to contribute many new organic measurements in the near future since most of the organic species do not have unique optical properties and are at very low concentrations. One exception is formaldehyde (CH2 O) and satellite measurements of this compound have been used elegantly to constrain the precursor isoprene emissions (Palmer et al. 2003). Some species, including CFCs-11, -12, -22, CCl4 and ethane, have been measured from a balloon using a Michelson interferometer to 3 km resolution, giving hope for the next generation of satellite measurements. Further candidates for future measurement by satellite are PAN and acetylene. It may also be possible to use typical emission ratios in conjunction with CO measurements to improve estimates in inaccessible regions. Satellite data drive weather forecasts which are increasingly likely to include ozone predictions in the future, as air quality is increasingly recognised as a meteorological hazard (Beer 2001). However, organic trace gas forecasts (e.g. Lawrence et al. 2003) are currently only of academic

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interest and available to the public only through a small number of internet sites (e.g. http://dionysos.mpch-mainz.mpg.de/∼lawrence/forecasts.html). Over the past hundred years global anthropogenic emissions of organic species have increased by an estimated factor of seven (van Aardenne et al. 2001), and in the future they can be expected to change still more as land use, vegetation patterns and industrial development alter rapidly (Andreae et al. 2002; Steiner et al. 2002). Since the 1980s the growth in surface emissions in North America and Europe has lessened, although in other areas such as the tropics strongly increasing trends are expected in line with economical development. Emissions from air traffic, which affects cirrus clouds and thereby climate, are also predicted to increase in the future (Seinfeld 1998; Travis et al. 2004). Control of anthropogenic organic emissions is politically difficult to realise both on regional (Grant et al. 1999) and on global scales (e.g. Montreal Protocol http://www.unep.org/ozone/montreal.shtml). To date most of our research on organics in the atmosphere has been passive. We have attempted to quantify emissions into the atmosphere and make deductions from there. On the other hand, deliberate emission of certain species has been proposed and practised. One proposal was the deliberate emission of hitherto unused compounds, of known lifetime, in specific amounts into the Earth’s atmosphere. Through regular measurements over a suitable time period, it would then be possible to determine accurately the global HO strength (Jöckel et al. 2003). Toxic compounds are also intentionally emitted into the environment as pesticides (Baker et al. 1996). Less than 0.1% of these applied pesticides reach their target raising important ethical questions (Pimentel 1995). It has been suggested that by actively injecting organics (ethane and propane) into the polar stratosphere, ozone depletions may be reduced (Cicerone et al. 1991). Follow-up research suggested, however, that this could have exactly the opposite effect (Elliot et al. 1994), which represents a stark warning against hasty experimentation. Recently it has been proposed to investigate the possibility of injecting sulphur or black carbon aerosol into the stratosphere as a method to attenuate incident radiation and hence engineer an atmospheric cooling to offset global warming effects (Crutzen 2006). We may also be tempted to take active measures if, in the distant future, the planet begins to cool. Humankind may well consider dosing the atmosphere with some of radiatively active organic gases to offset ice ages or to adjust atmospheres of other planets. There is clearly an important role for organic species in the atmosphere long into the future.

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Volatile Organic Compounds in the Atmosphere Edited by Ralf Koppmann Copyright © 2007 by Blackwell Publishing Ltd

Chapter 2

Anthropogenic VOCs Stefan Reimann and Alastair C. Lewis

2.1

Introduction

VOCs were first identified as being involved in photochemical smog formation in the 1950s (Haagen-Smit 1952), and since that time our knowledge of their behaviour in the environment has grown enormously. In this chapter an overview is given on the temporal evolution of anthropogenic VOC emissions in different parts of the world and the relative importance of anthropogenic source groups. This is followed by a discussion about the atmospheric distribution of the anthropogenic VOCs from polluted urban areas to the most remote parts of the globe. The chemical behaviour of, atmospheric VOCs is discussed in relation with the formation of ozone and organic aerosols. The final section of the chapter discusses the analytical methods for the detection of anthropogenic VOCs in the atmosphere. The emphasis of this chapter is on non-methane hydrocarbons (NMHCs) as oxygenated and halogenated hydrocarbons will be discussed separately in Chapters 4 and 5, respectively.

2.2

Sources of anthropogenic VOCs

The anthropogenic fraction of atmospheric VOCs is related to the unprecedented usage of fossil fuels for transport, the production of consumer goods and various industrial processes in the past centuries. Mainly three groups of anthropogenic VOCs can be distinguished: NMHCs, oxygenated hydrocarbons (OVOCs) and halogenated hydrocarbons (e.g. chlorofluorocarbons [CFCs] and hydrofluorocarbons [HFCs]). These substances are either directly emitted to the atmosphere or formed during combustion processes. The global emission of anthropogenic VOCs is estimated to amount to 186 Tg/year (EDGAR 2005). However, the majority of VOCs in the atmosphere are related to biogenic emissions, with estimated totals between 377 TgC/year (i.e. 442 Tg/year) (IPCC 2001) and 1 150 TgC/year (Guenther et al. 1995), which originate nearly exclusively from vegetation and small contributions from oceans and soils. The distinction between biogenic and anthropogenic VOCs in the atmosphere is far from straightforward because many VOC species are produced by both sources. Emissions of alkanes and alkenes, for example, are dominated by anthropogenic sources, but are also produced by soils, wetlands and oceans. On the other hand, OVOCs like acetone and methanol have dominant sources related to emissions from vegetation (Goldstein and Schade 2000;

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Jacob et al. 2002, 2005; Singh et al. 2004), but are also emitted by from combustion of fossil fuel and biofuel. Furthermore, OVOCs are also produced by the oxidation of hydrocarbons during atmospheric transport, which contributes considerably to their abundance at rural and remote sites (Goldstein and Schade 2000; Singh et al. 2004; Solberg et al. 1996). Related to the distinction between biogenic and anthropogenic VOCs in the atmosphere, the newly developed measurements of isotopes of VOCs could become an important analytical tool (Goldstein and Shaw 2003). To make things even more difficult, the definition of anthropogenic sources itself is delicate, as biomass burning, which could also be regarded as a natural process, has been dramatically enhanced by agricultural practices in the tropics. In this chapter the sources of VOCs will be discussed, using both emission inventories (bottom-up) and emissions derived from modelling of ambient measurements (top down). First, an overview on VOC emissions from different parts of the world and their temporal evolution will be given by using global and regional inventories. Then sources will be distinguished by emission activities and in relation with different groups of VOCs. Finally, methods for the modelling of emissions from measurements will be discussed.

2.2.1

Inventories for anthropogenic VOCs

The EDGAR inventory is a world-wide systematic approach to estimate emissions of anthropogenic greenhouse gases and air pollutants (including VOCs) down to a regional scale. It contributes to the Global Emissions Inventory Activity (GEIA) of IGBP/IGAC and has been used in the IPCC Assessments (IPCC 2001). Within EDGAR, sources are divided into 28 categories, which roughly can be aggregated to biofuel combustion, fossil fuel use and production, industrial processes, agricultural practices, biomass burning and waste treatment. Latest emissions are available for the year 2000 (EDGAR 32FT 2000) on a 1◦ × 1◦ grid (EDGAR 2005) (Figure 2.1). Global anthropogenic VOC emissions in the year 2000 were estimated to be 186 Tg/year, which represents a continuing increase in comparison with 1990 (153 Tg/year) and 1995 (160 Tg/year). The gap of over 20 Tg/year between 1995 and 2000 was predominantly caused by the biomass burning source in 2000 (49 Tg/year) being more than twice as high as in 1995 (22 Tg/year). This is due to a methodological change in the calculation of biomass burning emissions, using observations of burned area instead of burned area statistics, leading to enhanced emissions in Latin America and Africa. In Figure 2.2 the relative contributions of VOC emissions from aggregated activities are shown for the world, East Asia, the OECD (Organisation for Economic Co-operation and Development) and the United States. On a global average, the use of fossil fuels accounts for about 40% of the emitted VOCs. About 25% of the emissions are due to biomass burning and the rest is about equally divided between the use of biofuel and industrial processes. In East Asia (i.e. mainly China), biofuel combustion (e.g. from coal-burning cook stoves) and fossil fuel usage are dominating, with smaller emissions from industrial activities. In the OECD countries and in the United States, the partition between the sources is comparable, with about 50% of the VOC emissions caused by fossil fuel use, followed by a huge contribution from industrial processes and only minor sources from the remaining activities.

2–10 10–50 50–100 100–2000

Figure 2.1 Gridded VOC emissions from anthropogenic sources in the year 2000 from EDGAR 32FT2000 (With permission from EDGAR 2005). This image appears in full colour in the plate section that follows page 268 as Plate 1.

Unit: Gg NMVOC =0 0.0–0.1 0.1–1 1–2

Global total: 1.8e + 11 kg (min. = 60.9, max. = 1.1e + 009)

36

Volatile Organic Compounds in the Atmosphere

100%

80%

60%

40%

20%

0%

Global average

USA

OECD

East Asia

Waste management

2701

717

498

215

Biomass burning

48746

1686

163

257

Biofuel combustion

30013

748

288

6093

Industrial processes

26669

6650

5339

2566

Fossil fuel use, production

77428

9129

8040

5642

Figure 2.2 A comparison of aggregated activities related to VOC emissions (Gg/year) in the world (global average), United States, OECD countries and East Asia (EDGAR 2005).

Furthermore, emission inventories on a continental level have been developed in Europe and the United States. For Europe the CORINAIR-EMEP initiative provides yearly emission estimates, either down to a 50 × 50 km2 grid, as national totals, or by activity sectors (EMEP 2005; Vestreng 2003). For the United States the National Atmospheric Emission Inventory (NEI) (NEI 2005) from U.S. EPA estimates emissions on an annual basis since 1970 by defining three classes of sources: point sources (e.g. power plants), area sources (aggregates of small point sources) and mobile sources (vehicles and marine vessels). The ACESS project (Ace-Asia and Trace-P Modelling and Emission Support System; ACESS 2005) is a remarkable attempt to provide accurate emission estimates for the Asian continent. Emissions from this part of the globe are extremely difficult to assess because economical parameters are changing fast and a great variety of poorly defined sources exists (e.g. coal-burning cook stoves, biomass combustion). Emissions are either available for particular VOC species or on a 1◦ ×1◦ grid for the total VOCs. The total Asian anthropogenic emissions were estimated to be 52.2 Tg/year in the year 2000 (ACESS 2005; Streets et al. 2003) with residential combustion of coal and biofuels and transportation having the highest contribution. For China, Klimont et al. (2002) have estimated the VOC emissions in the year 2000 to be 15.6 Tg/year. Due to rapid growth in the use of personal transport and solvents in China, emissions are anticipated to grow fast; however, technical improvements (e.g. catalytic converters) are expected to mediate the increase (Klimont et al. 2002). This behaviour is in contrast to the development in the United States and Europe, where emissions have decreased for many years as a direct result of control strategies such as catalytic converters, better fuel efficiencies and an only moderate economic growth (Figure 2.3).

Anthropogenic VOCs

37

40

Tg/year

30 USA EU-15 China

20

10

0 1970

1980

1990

2000

2010

2020

Figure 2.3 Anthropogenic VOC emission estimates and projections from inventories in Europe (EU-15) (EMEP 2005; Vestreng 2003), United States (NEI 2005) and China (Klimont et al. 2002).

2.2.1.1

Regional and urban inventories

Numerous emission inventories have been developed in support of studies related to air quality, atmospheric chemistry and climate. In Europe examples for countries with a sophisticated emission inventory are Great Britain (National Atmospheric Emission Inventory, NAEI) (NAEI 2005) and Germany (Friedrich et al. 2002). In Great Britain emissions have been regularly estimated for many years on a very detailed level in terms of VOC species, spatial resolution and anthropogenic activities (Figure 2.4a). Lately, the German inventory has substantially improved by incorporating measurements at specific sources (e.g. validating the exhaust source with chassis dynamometer tests) (Schmitz et al. 2000). Serious efforts have been made within urban environments around the world to develop emission inventories for VOCs and other compounds related to tropospheric ozone formation. As an example for an inventory within a developing country, Figure 2.4b shows VOC emissions from the city of Delhi (India) (Gurjar et al. 2004). Figure 2.4 shows the dominant effect of transportation in Delhi in comparison with the moderate fraction related to road transport in Great Britain. Furthermore, total emissions in Great Britain decline since the early 1990s, whereas emissions in Delhi are still increasing at the end of the 1990s, although to a possibly smaller extent in the last years.

2.2.2

Important sources of anthropogenic VOCs

In this section the source activities which lead to emissions of anthropogenic VOCs will be discussed, using grouped activities and quantitative emissions for the year 2000 from EDGAR (2005).

2.2.2.1

Fossil fuel use and production

Largely three source categories can be distinguished within this group: mobile sources, stationary sources and sources due to production, storage and delivery of fossil fuel products. Emissions from this activity are estimated to sum up to a global total of 78 Tg/year,

38

Volatile Organic Compounds in the Atmosphere

3

(a)

Road transport Other transport

Tg/year

2

Extr and Distr fossil fuels Solvents 1

Processes Other

0 1970

1975

1980

1985

1990

1995

2000

(b) 250

Gg/year

200 Industry 150

Domestic Transport

100

Power plants

50 0 1990

1995

2000

Figure 2.4 VOC emissions from main source activities in (a) Great Britain (NAEI 2005) and (b) city of Delhi. Adapted from Gurjar (2004).

predominately due to road transport and oil production. For road transport, areas of highest emissions are located in the United States, Europe, Latin America and South East Asia, whereas highest shares of emissions from oil production can be found in regions with biggest oil drilling activities (i.e. Middle East, Latin America). Mobile sources can be further divided into emissions from the exhaust and fugitive emissions by evaporation. As VOCs from fossil-fuel-driven vehicles globally represent the most important anthropogenic activity, bottom-up inventories using dynamometer test measurements under laboratory conditions have a long history. With those experiments it is not only possible to assess emissions of VOCs from tailpipe exhausts (Heeb et al. 2000a; Schmitz et al. 2000; Siegl et al. 1999) but also to estimate non-tailpipe, fugitive emissions by evaporation (Pierson et al. 1999). According to the valid vehicle legislation, exhaust emissions are measured within driving cycles such as the US FTP-75 and the European driving cycle (EDC). In addition, socalled ‘real-world’ cycles are applied as well on chassis dynamometers. These cycles better represent transient driving at all relevant traffic situations, for example, urban stop-andgo, extended highway driving and driving under cold start conditions. The overall vehicle emissions are then extrapolated by correctly weighting emissions for all relevant driving patterns and vehicle classes. Technical progress like the introduction of catalytic converter systems and more energy-efficient motor systems led to lower emissions per driven distance

Anthropogenic VOCs

39

100% VOCs Benzene 80%

60%

40%

20%

0%

1994

1995

1996

1997

1998

1999

2000

2001

Figure 2.5 Relative decrease of average light-duty car exhaust emissions for VOCs and benzene in summers between 1994 and 2001 in California, with emissions in 1994 corresponding to 100%. Adapted from Kean et al. (2002).

for newer car models (Heeb et al. 2000b; Kean et al. 2002) (Figure 2.5). In addition, fuel composition has been changed over the years for both technical and health reasons. For example, the reduction of carcinogenic benzene in fuel has led to an over-proportional decline of benzene emissions from car exhausts in comparison with the remaining VOCs (Figure 2.5). However, as a noble metal based three-way catalyst also produces benzene from other aromatic compounds (Bruehlmann et al. 2005), the decrease of benzene levels in the fuel is only partially represented in the exhaust emissions (Heeb et al. 2002). Furthermore, time-resolved emission measurements on chassis dynamometers showed that in the wake of emission reductions by catalytic converters ‘cold start emissions’ are becoming more and more important (Baugh et al. 1987; Heeb et al. 2003; Weilenmann et al. 2005). The majority of the VOC emissions of today’s gasoline-fuelled vehicles are released during the first few minutes after the engine start. These so-called ‘cold start emissions’ correspond to the time needed to reach the catalyst light-off temperature. Cold start emissions are relevant especially in urban areas, where short-range journeys represent a considerable part of driven distance. Furthermore, chassis dynamometer tests and sealed housings were used to assess the fugitive emissions from passenger cars (Pierson et al. 1999). Fugitive emission rates increase with both, higher ambient temperature and higher vapour pressure of the fuel contents. This behaviour was confirmed by many studies in the warm season, where significant fugitive emissions from car tanks were found (Borbon et al. 2003b; Harley et al. 1992; Kourtidis et al. 1999; So and Wang 2004). For assessing the real-world picture of VOC emissions due to road transport, tunnel measurements are a common method, as they give a representative image of the average emissions of the actual car fleet of a certain area. Furthermore, they can be applied to validate emission inventories used for air pollution control models (Staehelin et al. 1998) or to check the effect of changing fuel compositions (Kirchstetter et al. 1999).

40

Volatile Organic Compounds in the Atmosphere

Table 2.1 VOC measurements in tunnels used to assess VOC emissions from mobile sources Year (reference) 1993 (Staehelin et al. 1998) 2003 (Stemmler et al. 2005) 1997 (Schmid et al. 2001) 1998 (Kurtenbach et al. 2002) 1996 (Touaty and Bonsang 2000) 1990 (Haszpra and Szilagyi 1994) 1994/97 (Kirchstetter et al. 1999) 1999 (Kean et al. 2001) 1992 (Gertler et al. 1996) 2000 (McGaughey et al. 2004) 1994 (Duffy and Nelson 1996) 2000 (Hwa et al. 2002) 2000 (Na et al. 2002)

Tunnel location

VOC species

Gubrist, Zurich (Switzerland) Gubrist, Zurich (Switzerland) Tauern (Austria) Wuppertal (Germany) Thiais, Paris (France) Budapest (Hungary) San Francisco Bay (USA) San Francisco Bay (USA) Fort McHenry/Tuscarora (USA) Houston (USA) Sydney Harbour (Australia) Taipei (Taiwan) Seoul (South Korea)

−C10 , OVOCs C2 − −C8 C4 − −C10 , OVOCs C6 − −C10 , OVOCs C2 − −C6 C2 − −C10 C2 − −C9 , OVOCs C2 − OVOCs −C10 C2 − −C8 , MTBE C5 − −C10 C2 − −C9 C2 − −C9 C2 −

In Table 2.1 important tunnel studies from different regions of the world are summarised. Interestingly, some of the tunnels were re-analysed after a certain time lag, which provides a valuable approach to study the success of emission reduction measures. Thus, emission reductions for VOCs in the range of 80–90% were found in Austria and Switzerland, when VOCs were re-measured in the same tunnels (Schmid et al. 2001; Stemmler et al. 2005). Furthermore, emission estimates for the Swiss tunnel were also in compliance with the historical and actual bottom-up inventories for mobile VOC sources for Switzerland. This indicates that the development of the VOC emissions during the last decade is well understood on the basis of fleet composition and dynamometric test measurements. Contrary to these results, calculated emissions from a recent tunnel study in Houston were higher than those from former studies (McGaughey et al. 2004). This atypical result was explained by high ambient temperature during the most recent study, leading to higher fugitive emissions in the tunnel. In some of the tunnel studies also OVOCs were measured. Relative to the NMHC emissions, OVOC were found to be emitted in slightly lower, but still considerable, amounts (Schmid et al. 2001; Staehelin et al. 1998). In a tunnel in Houston (USA), methyl-t butylether (MTBE) was found to be one of the most abundant VOC species in tunnel air, which was related to fugitive emissions (McGaughey et al. 2004). Kean et al. (2001) found formaldehyde to be the most abundant OVOC (excluding MTBE) (accounting for 45% of the OVOCs) followed by acetaldehyde, aromatic aldehydes and acetone. They also detected a decline in emissions for the aldehydes, in the range of 45–70% between 1994 and 1999 from measurements in the same tunnel. Stationary emissions from the use of fossil fuel are due to industrial applications (e.g. refineries and coke ovens), residential heating in the cold season or cooking. Globally, emissions are relatively small compared to mobile sources. However, for certain regions (e.g. China) stationary sources can be a substantial part of the emissions. One of these sources is the extensive use of coal and biofuels in stoves and cookers in the residential

Anthropogenic VOCs

41

sector, which is estimated to contribute to about 30% of total Chinese anthropogenic VOC emissions in the year 2000 (Klimont et al. 2002). A meticulous study of emissions from different types of stoves and biofuel showed that considerable differences exist for emissions of VOCs per produced energy unit (Tsai et al. 2003). Emissions related to production, storage and delivery of fossil fuels predominately occur in those regions where extensive fossil fuel drilling activities exist. However, fugitive emissions can also occur from the transport and distribution of the fuel, such as ships, road tankers and fuel stations. Especially, emissions from fuel stations can be substantial on a local or regional scale. This is not only of concern related to the formation of tropospheric ozone, but also to health issues, as considerable emissions of toxic compounds (i.e. benzene and MTBE) can occur. Natural gas emissions from the delivery infrastructure (pipelines, installations) have been found to be important in urban areas. Their source strength is temperature dependent (Borbon et al. 2003b) and leads to elevated concentrations primarily of ethane and propane (Derwent et al. 1995).

2.2.2.2

Industrial processes

While methods to calculate VOC emissions from fossil fuel use and production are well established, the knowledge of emissions from industrial processes is less certain. The global emissions of 27 Tg/year (EDGAR 2005) are mainly due to the use of VOCs as organic solvents with different physical and chemical properties (i.e. alkanes, ketones, aldehydes, esters and halogenated hydrocarbons). Solvents are used in many consumer goods, such as paints, adhesives and inks, for which the evaporation of the solvent is essential after usage. This leads to elevated emissions at higher temperature during the warm season of the year (Borbon et al. 2003b). Other, more industrial uses encompass the manufacture of pharmaceutics, the cleaning of metal surfaces, the extraction of oil seeds or printing. Nowadays, many of these industries have reduced their emissions in developed countries by either recycling VOCs or thermally destroying them. However, emissions in other parts of the world may still increase. Thus, Chinese VOC emissions from solvents and paints are expected to be nearly five times higher in 2020, compared to 1990 (Klimont et al. 2002).

2.2.2.3

Biofuel combustion

Biofuel combustion is a primary source of energy in developing countries. The EDGAR inventory for the year 2000 estimates highest regional emissions in Asia (13 Tg/year) and Africa (8 Tg/year) with a global total of 30 Tg/year (EDGAR 2005). The overwhelming part of the emissions derives from heating and cooking in the residential sector. In South America, additionally, a considerable part of VOC emissions from the transport sector is assigned to the biofuel source. This is due to Brazil’s usage of ethanol (derived from plant materials) as vehicle fuel.

2.2.2.4

Biomass burning

Especially in the tropics biomass burning is a common practice in agriculture. In the year 2000 global biomass burning emissions of VOCs was estimated to be 49 Tg/year (EDGAR 2005) with highest contributions from Latin America, Africa and South East Asia.

42

Volatile Organic Compounds in the Atmosphere

Andreae and Merlet (2001) reviewed the literature for gaseous and aerosol emissions of biomass burning. Their global VOC emissions attributed to biomass burning is 60 Tg/year, which is divided into emissions from savannah and grassland (26 Tg/year), tropical forests (22 Tg/year) and extra tropical forests (12 Tg/year). They attributed around 50% of the emissions to oxygenated species like alcohols, ketones, aldehydes and acids.

2.2.2.5

Waste management

VOC emissions due to waste management are low in comparison with other sectors and are estimated to account for about 1–2% of the total anthropogenic VOC sources (EDGAR 2005). Emissions arise from waste incinerators due to incomplete burning and from waste dumping sites, either directly (e.g. by evaporation of deposited solvents) or as organic decay products.

2.2.3 2.2.3.1

Sources of VOC classes Alkanes

Due to its low reactivity and considerable global sources of about 10–13 Tg/year (Gupta et al. 1998; Rudolph 1995), ethane is the most abundant alkane in the atmosphere. The main sources are emissions from exploitation and distribution of natural gas and from biomass burning (Rudolph 1995). Minor sources include emissions from engine exhaust and biogenic emissions from soils and wetlands. Propane as the next higher alkane in the homologous series has main anthropogenic sources related to the distribution of natural gas −C5 -alkanes are primarily emitted from evaporation of and petrochemical industries. C4 − fossil fuel, which is stored in tanks and from the exhaust. C6 -alkanes and higher homologous VOCs are mostly emitted as a consequence of their usage as solvents and from fossil fuel evaporation. However, as the vapour pressure gets lower with the increasing chain length of the VOC, the relative importance of the evaporation becomes smaller for higher molecular VOCs. On the contrary, spillage of fossil fuel becomes more and more significant, especially in countries where measures for curbing refuelling losses are not in place.

2.2.3.2

Alkenes

Fossil fuels contain only small amounts of alkenes. Therefore, alkenes are emitted mainly from the vehicle exhaust (due to incomplete combustion of the fuel), from biofuel combustion and biomass burning. Ethene and propene are the most abundant species in polluted air masses. However, their concentrations decline rapidly in more remote areas because of their high reactivity in the atmosphere. Higher molecular alkenes are emitted in lower amounts, but can be important as precursors of tropospheric ozone in urban environments. Furthermore, alkadienes are emitted as a by-product of incomplete combustion processes. Examples are the carcinogenic 1,3-butadiene and isoprene. Interestingly, the latter was long believed to be exclusively of biogenic origin but its correlation with other products from fossil fuel combustion (Borbon et al. 2001; Reimann et al. 2000) suggests a small additional anthropogenic source.

Anthropogenic VOCs

2.2.3.3

43

Alkynes

Ethyne (acetylene) is the dominant alkyne in the atmosphere. In urban areas it is emitted from the incomplete burning of fossil fuel. In other areas (e.g. the tropics) emissions from biomass burning and biofuel combustion are most important (Andreae and Merlet 2001). Furthermore, small amounts of propyne and butyne are emitted by combustion processes and can be measured in polluted environments.

2.2.3.4

Aromatics

Aromatic VOCs are important components of fossil fuels and are predominantly emitted by vehicle exhausts, from fuel evaporation and spillage. In addition, toluene and xylenes also have significant industrial emissions because they are used as solvents. The most abundant species in fossil fuels and consequently in urban air masses are benzene, toluene, ethyl benzene, xylenes, styrene and trimethyl benzenes. As benzene has carcinogenic properties it has been replaced in fossil fuels by other aromatics or MTBE. Benzene is also produced during the combustion of biofuel (Barrefors and Petersson 1995), which can be important, for example, in regions with a high ratio of residential wood burning.

2.2.3.5

OVOCs

Oxygenated VOCs are emitted from both anthropogenic and biogenic sources, although on the global scale their predominant sources are emissions from vegetation and biomass burning. A significant fraction of OVOCs in the atmosphere are produced by the photochemical conversion of NMHCs to OVOCs (Jacob et al. 2002; Millet et al. 2004; Singh et al. 2000, 2004; Solberg et al. 1996). This secondary formation results in concentrations of OVOCs being normally higher than those of NMHCs in remote areas (Millet et al. 2004; Singh et al. 2004). Direct anthropogenic emissions of OVOCs are related to the incomplete combustion of fossil fuel from mobile sources and the use of OVOCs as solvents. Highest contributions from vehicular sources are from formaldehyde, acetaldehyde, Oxygenated aromatics and acetone (Kean et al. 2001). In addition, the fuel additive MTBE is emitted by evaporation and the exhaust (Kawamoto et al. 2003). However, due to possible detrimental effects on health and environment its use in gasoline has already been phased-out in California and probably will be on a global scale in the near future. After the phase-out, average concentrations of MTBE in California declined from over 1 ppb to 0.15 ppb (CARB 2005).

2.2.4

VOC emission estimations from atmospheric measurements

The aim of many atmospheric VOC measurements has been the assessment of VOC sources and their accurate coverage in emission models. This so-called top-down approach uses the temporal and spatial behaviour of ambient concentrations at given receptors (measurement sites) to infer emissions from a certain region. One of the most straightforward methods is the tracer-ratio technique, where measurements of a substance with unknown source strength are related to concurrent measurements of a substance with known emissions (a priori information). This technique is often

44

Volatile Organic Compounds in the Atmosphere

used in urban studies, where the chemical decay during transport has a negligible influence on measured concentrations (due to the short distance between the source and the receptor). Examples are the estimation of VOC emissions in Great Britain (Derwent et al. 2000), in Germany (Augsburg (Klemp et al. 2002; Mannschreck et al. 2002; Slemr et al. 2002a), Berlin (Winkler et al. 2002)), in Italy (Lombardy (Dommen et al. 2003)), in China (Hong Kong (Guo et al. 2004a)) and in Mexico (Mexico City (Arriaga-Colina et al. 2004)). Resulting emissions from the top-down approach can differ considerably from those of emission models. Thus, discrepancies between calculated and measured emissions in Augsburg were observed related to the relative contribution of source categories in different seasons (Mollmann-Coers et al. 2002). An illustration for the application in a remote atmosphere is the estimation of global emissions of relatively stable OVOCs from biomass combustion by measurements over the Pacific Ocean (Singh et al. 2004). For the determination of the relative contribution of different sources to measure atmospheric VOCs, the following equation is often taken as the basis: X (n, m) = S(n, p) × Q(p, m), X represents the n × m matrix with a number of n measurements of the concentrations of m species. S is the matrix of the source contributions (n × p matrix) and Q is the matrix of the source profiles (p × m matrix, with p sources). There are two principal ways to solve this equation. The first is the Chemical Mass Balance (CMB), where it is assumed that the source profiles (Q) are known and the equation can be solved with a least-square fit. In the second approach, both S and Q are assumed to be unknown and the equation is solved with multi-variate statistical methods. Both methods have the principal limitation that the mass conservation is implicitly presupposed, which due to atmospheric reactions is only partly fulfilled. For the CMB model many studies have been performed in urban environments. Examples for the United States (Watson et al. 2001) are studies in Texas (Fujita 2001), Chicago (Scheff and Wadden 1993) and Michigan (Scheff et al. 1996). In Europe emissions have been estimated, for example, in Birmingham (Great Britain) (Hopkins et al. 2005) and in Helsinki (Finland) (Hellen et al. 2003). Lately, Zhao et al. (2004) developed the CMB method further by introducing measures to account for meteorological factors and human behaviour (i.e. weekdays/weekends). For the multivariate methods the Principal Component Analysis (PCA) is a widely used tool for analysing emission sources with ambient measurements (Buhr et al. 1992). With the help of correlated variables (e.g. concentrations, meteorological data) a smaller number of independent factors is identified (i.e. principal components, PCs) which can explain the variance in the data. The number of extracted PCs corresponds to the number of sources. This approach has been successfully used in many urban environments, such as London (Great Britain) (Derwent et al. 1995), Lille (France) (Borbon et al. 2003a), Zurich (Switzerland) (Staehelin et al. 2001) and Hong Kong (China) (Guo et al. 2004b). An example for the application of the PCA model in a more remote environment is a study on source contribution at a rural site in China (Guo et al. 2004c). Another multi-variate method is the Positive Matrix Factorisation (PMF), which unlike the PCA, does not rely on the information from the correlation matrix but utilises a

Anthropogenic VOCs

45

weighted least-squares minimisation scheme. Examples of source apportionment studies are from Santiago de Chile (Jorquera and Rappengluck 2004) or Houston (Kim et al. 2005; Zhao et al. 2004). An alternative approach for the detection of different sources, which contribute to ambient concentrations at a given receptor, was developed by Jobson et al. (2004) by adopting earlier work related to VOC reactivities (Jobson et al. 1998). The method is based on the fact that in an urban environment a significant part of VOC concentration variability is due to variations in emissions. Applying this method, traces of industrial emissions have been detected at the side of the dominant vehicle exhaust source (Jobson et al. 2004). Furthermore, chemical transport models (CTMs) are used for the estimation of VOC emissions. Thereby, measured concentrations are compared with those resulting from the CTM model, when an emission inventory is used for the initialisation of the model and the chemistry and transport processes are accounted for. CTMs have, for example, been used successfully to validate VOC emissions from Europe by modelling measurements in Scandinavia (Hov et al. 1997) or on a smaller spatial scale to model emissions from urban areas like Paris (France) (Vautard et al. 2003) and Mexico City (Mexico) (West et al. 2004).

2.3

Atmospheric distribution of VOCs

The spatial and temporal distribution of VOCs in the atmosphere depends on the location of their sources, their reactivities and regional and global transport phenomena. In this section, a short historic introduction will be followed by a discussion of observations, ranging from urban centres to global background. Furthermore, general features like diurnal and annual cycles, as well as the spatial and temporal behaviour of VOC concentrations under different conditions will be discussed.

2.3.1

Historic VOC concentrations

The analytic methods for measuring VOCs in ambient air have been developed in parallel with the rapid increase in fossil fuel usage. First measurements were pioneered in the Los Angeles area, where air pollution problems became overwhelming in the 1950s. The most important analytical tool ever since has been gas chromatography (GC) with prior extraction from air onto sampling traps and subsequent analysis using detection by thermionic conductivity (TCD), by flame ionisation (FID) or mass spectrometry (MS). Probably the first ambient VOC measurement ever was made in 1956 (Eggertsen and Nelsen 1958), when VOCs from 6.4 l of Los Angeles air were analysed by a GC equipped with a TCD (Figure 2.6). Concentrations were extraordinarily high compared to today’s values (e.g. 180 ppb of ethyne, 150 ppb of propene and 130 ppb of isopentane). In fact, these high concentrations were required for VOCs to be detectable with the chosen method: ‘The threshold of detection is 0.05 p.p.m. or better, which brings the technique well within the concentration range of interest in urban air pollution studies’ (Eggertsen and Nelsen 1958). First measurements of VOCs using GC–MS followed (Quiram and Biller 1958), with an indicative concentration of 110 ppb of total VOCs. Further studies were performed in the same area, with first measurements of toluene in the atmosphere (i.e. 60 ppb) in 1958 on a

45

Charging 6.4 l

30

0°C

3

Ethane

Trapping column (dimethylsulfolane on firebrick)

−183°C

0.0

0.1

0.2

0.3

0.4

0

‘Air’

10

20

Isopentane

30

n-Pentane

Separating column (dimethylsulfolane on firebrick)

n-Butane Acetylene

Isobutane

Propylene

Ethylene

Propane + N 2O

40

Propyne + I-pentene

Figure 2.6 The first published chromatogram of ambient VOC measurements showing light hydrocarbons measured in 1956 in Los Angeles (Eggertsen and Nelsen 1958). With permission from the American Chemical Society.

Peak height (mV)

Anthropogenic VOCs

47

day where ‘the smog concentration was officially reported as moderate to heavy’ (Farrington et al. 1959). From the very beginning of these measurements vehicular traffic was suggested to be the predominant source of atmospheric VOCs, as concentration patterns were identical to those found in road tunnels (Eggertsen and Nelsen 1958; Farrington et al. 1959). Additional sources were proposed to contribute to the ambient concentrations, such as losses from gasoline evaporation, natural gas distribution, solvents and oil fields (Altshuller et al. 1971). First diurnal cycles of VOCs were shown by Gordon et al. (1968) and by Altshuller et al. (1971), with a morning maximum related to traffic, followed by a decline during the day hours and an increase in the late afternoon. This was also confirmed by Lonneman and Bufalini (1978), who found that vehicular emissions were dominant on weekday mornings but were normally smaller during the afternoon and on weekends. Simultaneously discussions about the limitations of air pollution policy measures already started and it was speculated that the complete control of automotive emissions, other combustion sources and organic solvents in the Los Angeles basin still is not likely to reduce hydrocarbons to the desired level, if most of these paraffinic hydrocarbons were associated with natural gas leakage and petroleum gas leakage (Altshuller et al. 1971). In order to follow the development of anthropogenic VOC emissions in the past 50 years it is illustrative to compare measurements in California starting in the 1950s (Singh and Zimmerman 1992) with the current average yearly concentrations in the same area (CARB 2005) (Figure 2.7). In the beginning, concentrations of benzene and toluene were about

60

50 Benzene Toluene

ppb

40

30

20

10

0 1955

1965

1975

1985

1995

2005

Year

Figure 2.7 Benzene and toluene concentrations in California in the past 50 years (CARB 2005; Farrington et al. 1959; Sexton and Westberg 1984; Singh and Zimmerman 1992).

48

Volatile Organic Compounds in the Atmosphere

40 and 60 ppb, respectively. Since then, levels have declined dramatically and average values are below 1 or 2 ppb for benzene and toluene, respectively. This indicates the ongoing success of the technical solutions such as catalytic converters, which counterbalances the ever increasing vehicular traffic.

2.3.2

Common features of VOCs in polluted areas

In urban atmospheres VOCs exhibit certain recurrent patterns, which are dependent on human behaviour, the geographical location, meteorological parameters and the timedependent contribution of different sources. A typical diurnal cycle for an urban atmosphere is illustrated by measurements from the city centre of Zurich (Figure 2.8). As in many major cities, VOCs in Zurich are emitted predominantly by vehicular traffic. Concentrations in the early morning hours are low, as emissions from the previous day have been diluted or degraded. After 6 am commuter traffic starts and concentrations rise sharply. From about 10 am onwards, concentrations begin to decline steadily until about 4 pm. This behaviour is not only caused by the smaller amount of traffic but also by an enhanced photochemical sink and increasing thermal convection during the daytime hours. This leads to a dilution of the polluted city air with cleaner air masses from above the convective boundary layer. During the evening hours commuter traffic leads again to increased concentrations, which especially on warm, calm days in summer coincides with a stratification of the air, caused by the faster cooling of the atmosphere near the surface relatively to the warm air above. Concentrations on a day-to-day basis are additionally influenced by meteorological phenomena. Generally, the horizontal wind velocity is the most important factor, leading to an efficient dilution of local and regional emissions. As an example, VOCs concentrations from Zurich are shown in Figure 2.9, where stormy conditions (7–9 February 2004) led to a strong decline of measured VOC concentrations. Thereby, the wind not only advects more 2.5

2.0 C4-alkanes C5-alkanes C2-benzenes Benzene Toluene

ppb

1.5

1.0

0.5

0 0

6 AM

12 PM

6 PM

0

Figure 2.8 Average diurnal cycles of C4 -alkanes (n-butane and isobutane), C5 -alkanes (n-pentane and isopentane), C2 -benzenes (ethyl benzene, m-, p-, o-xylene), benzene and toluene in Zurich in the year 2004.

Anthropogenic VOCs

49

9.0

6 Benzene Isopentane Wind velocity

7.5

4

6.0

3

4.5

2

3.0

1

1.5

0

m/s

ppb

5

0.0 4

5

6

7

8

9

10

11

Date in February 2004 Figure 2.9 Concentrations of benzene and isopentane in Zurich (February 2004) showing a negative correlation with the wind velocity.

6 Benzene Toluene C2-benzenes

ppb

4

2

0 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 Year

Figure 2.10 Monthly mean concentrations (1993–2004) of aromatic VOCs at a suburban site (Dubendorf, Switzerland).

pristine air masses to the city but emissions also get transported faster into the surrounding areas. However, this can only conditionally be seen as a general feature, as point sources in the pathway of the advected air masses could also lead to a concentration increase. The seasonal cycle of VOCs is characterised by higher concentrations in winter than in summer. As an example concentrations of aromatic VOCs from Duebendorf (in the surrounding of Zurich, Switzerland) between 1993 and 2004 are shown in Figure 2.10. This recurrent annual cycle is caused by a combination of the generally lower chemical reactivity of the atmosphere in the cold season and higher emissions from fossil-fuel-driven vehicles (‘cold start’, see above). Furthermore, especially in European winter, long-lasting periods

50

Volatile Organic Compounds in the Atmosphere

with constant atmospheric conditions can occur, where emissions are trapped below stable stratified clouds. In addition, Figure 2.10 illustrates the declining concentrations of aromatic VOCs within the last years in Western Europe. This feature is concurrent with the development in the United States, already discussed for Figure 2.7.

2.3.3

VOCs in urban areas

In this section an overview on important findings from campaigns in urban environments in different parts of the world is given. Concentrations of anthropogenic VOCs from selected cities show a common pattern with concentrations being lower in Western Europe and Northern America compared to the rest of the world (Table 2.2). In fact, actual Table 2.2 Overview of recent VOC measurements in urban areas from different parts of the world. Houston (Jobson et al. 2004), London (Derwent et al. 2000), Athens (Bakeas et al. 2003; Rappenglück et al. 1998), Chinese cities (Barletta et al. 2005), Hong Kong (Guo et al. 2004b), Porto Allegre (Grosjean et al. 1998), Karachi (Barletta et al. 2002) Houston,a USA 2000

London, GB 1996

Ethane n-Propane n-Butane Isobutane n-Pentane Isopentane n-Hexane

5.90 3.15 1.11 1.21 0.37 0.91–1.23 0.29–0.41

3.4–3.9 2.3–2.6 3.2–4.9 0.8–2.6 0.7–1.0 2.4–4.1 0.2–0.4

Ethene Propene S-butene 1,3-Butadiene S-cis/trans-pentene Isoprene

1.83 0.45 0.14 0.05–0.09 0.02 0.11–0.18

3.0–3.5 1.8–2.1 0.5–0.8 0.2–0.4 0.3–0.4 0.1

2.1–34.8 0.2–8.2 0.2–12.5 0.02–2.5 0.04–14.7 0.04–1.7

0.41

2.9–7.1

2.9–58.3

Benzene Toluene Ethyl benzene m/p-Xylenes o-Xylene

0.32–0.34 0.40–0.42 0.06 0.16 0.06

1.2–1.9 2.7–3.6 0.5–0.7 1.5–2.1 0.7–0.8

Methanol Formaldehyde Acetone Acetaldehyde

3.90 3.90 3.25–3.70 1.94–2.47

Ethyne

Athens, Greece 1994, 2000

1.6–12.4b 1.7–14.2b 3.2–26.3 0.6–4.1

2.5–11.7 6.7–21.2 1.3–4.0 3.2–11.1 1.5–5.5 10.7–17.2 6.5–13.7b 12.3–15.1

Notes: Concentrations are in ppb; S- = sum of. a Concentration compromised by other compounds. b Median of concentrations from different locations.

43 cities, China 2001

3.7–17.0 1.5–20.8 0.6–14.5 0.4–4.6 0.2–7.7 0.3–18.8 0.1–3.6

0.7–10.4 0.4–11.2 0.1–2.7 0.4–15.3 0.1–6.0

Hong Kong, Porto China Allegre, 2001 Brazil 1996

3.6–4.9 5.6–8.8 2.8–3.6 1.5–2.2 3.1–4.7

1.3–2.0 1.8–2.6 0.3–0.4 0.8–1.3 2.1–2.5 13.1–13.5 1.2–1.3 1.6–1.8 0.5–0.6

Karachi, Pakistan 1998/99

2.4 46.1 8.3 4.8 5.9 10.6 6.4

93.0 41.0 19.8 11.0 13.4 12.1 7.5

42.4 13.8 12.8 1.7 5.3 0.2

19.0 5.5 2.8 0.8 0.4 0.8

65.8

18.0

6.5 9.7 2.7 5.4 2.1

5.2 7.1 1.1 2.1 1.0

4.4 2.4

Anthropogenic VOCs

51

concentrations in many cities in developing countries are comparable to those from the first VOC measurements in Los Angeles almost 50 years ago. Partly, this is caused by higher evaporative emissions, especially in tropical regions, but mainly by less advanced technical solutions for efficient emission reductions. In North America the early studies in Los Angeles (Section 2.3.1) were followed by campaigns in practically all major cities. A summary of the work of the late 1970s is given by Sexton and Westberg (1984), where data from ambient air monitoring programmes in seven urban centres (Houston, Philadelphia, Baltimore, Washington, Newark, Boston, Milwaukee) and six rural areas are compiled. VOC compositions in cities were found to be more or less identical and the average VOC concentrations in rural areas were at least an order of magnitude below mean urban concentrations. More than a decade later Solomon et al. (2000) provided an overview of ozone-related studies in Northern America and Europe under a wide range of geographical and climatological conditions. Apart from Los Angeles, especially Houston (Texas) has been a major point of interest because of the influence of its large industrial facilities in the vicinity of the city. After Siddiqi and Worley (1977) first addressed the special conditions in Houston, it was recently found that several VOCs like short-chained alkenes and alkanes were systematically influenced by petrochemical industry emissions (Berkowitz et al. 2004; Kleinman et al. 2002). In Europe atmospheric VOC measurements were lacking some years behind the United States. One of the first measurements was performed in 1969 by Grob and Grob (1971), who measured 51 ppb of benzene and 39 ppb of toluene in the air of Zurich. Vehicular emissions are unambiguously detected as the main source, followed by evaporation of fuel (in summer), stationary combustion (in winter) and natural gas emissions. VOC concentrations in North Western European cities are comparable to those in Northern America (Derwent et al. 1995, 2000; Vautard et al. 2003; Winkler et al. 2002). On the other hand, cities in Southern Europe like Athens (Greece) (Rappenglück et al. 1998) and Rome (Italy) (Brocco et al. 1997) endure notorious air pollution problems during the warm summer days. In Athens, for example, pollution is enhanced by a local land-sea breeze, which exports air from the city in the morning and returns it in the afternoon (Flocas et al. 2003; Gusten et al. 1988; Rappenglück et al. 1998). Remarkably, maximum concentrations during stagnant periods reached 61 ppb for benzene and 103 ppb for toluene (Rappenglück et al. 1998) – which, once again, is comparable to values in the 1960s in Los Angeles. Vehicle emissions, fuel evaporation and paint solvents were found to be the main sources in Athens (Moschonas and Glavas 1996). In Italy several campaigns have been performed in the heavily industrialised Po Basin between 1998 and 2003 (Northern Italy) (Dommen et al. 2002; Duane et al. 2002; Neftel et al. 2002; Prevot et al. 1997; Steinbacher et al. 2005a, 2005b). Concentrations were influenced by both vehicle emissions and industrial solvents. A pronounced difference in the ratio of toluene to benzene was detected during weekends and during main holiday periods, indicating a much reduced amount of industrial emissions during these periods (Steinbacher et al. 2005b). In Japan and Australia measurements have been performed in several major urban agglomerations (e.g. Yamamoto et al. 2000). Resulting concentrations are comparable to those in Northern America, as reducing emissions, for example, by using catalytic converter systems, is a common practice in these countries. Diurnal variations were found to

52

Volatile Organic Compounds in the Atmosphere

be consistent with traffic activities and aromatic VOC concentration was in the lower ppb range. In Africa, South America and Asia the development of so-called megacities creates a big challenge for air pollution control policy. Funding for abatement measures is limited, population and traffic is growing and the inefficient traffic flow leads to a vicious circle with even higher air pollution (Gurjar and Lelieveld 2005). In Africa measurements are restricted to cities in the North of the continent (Abu-Allaban et al. 2002; Cecinato et al. 2002; Doskey et al. 1999). Concentrations are generally higher than in Northern America (e.g. benzene: 8–43 ppb, toluene: 17–111 ppb and n-butane: 71–174 ppb (Abu-Allaban et al. 2002)). In Central America, Mexico City endures enormous air pollution problems because its situation on a high altitude basin leads to a combination of trapped emissions and more efficient ozone formation. Average total VOC concentrations in the morning have been measured to be as high as 1.1–6.7 ppmC (Arriaga-Colina et al. 2004). Liquefied petroleum gas (LPG) represents 16% of the total fuel consumption in Mexico City, which effects in propane and butanes (the main compounds of LPG) to constitute 29% of the atmospheric VOCs (Gasca et al. 2004). In Southern America, Brazil is another focus of interest because ethanol is used as vehicle fuel and accordingly was found to be one of the most abundant VOCs in urban air (Grosjean et al. 1998, 2002). In Asia measurement campaigns have been performed in various urban areas (e.g. Seoul, South Korea (Na et al. 2003); Bombay, India (Rao et al. 1997) and Karachi, Pakistan (Barletta et al. 2002)). In most of the Asian cities indications for considerable solvent emissions have been found in addition to the dominant traffic source (Na et al. 2003; Rao et al. 1997). Concentrations of total VOCs are in the range of several hundreds of ppbs, indicating the need for more efficient emission control. In China the unconstrained growth of the transport sector and the lack of initiatives for measures to reduce emissions are the basis for considerable air pollution problems (Fu 2001). Roadside measurements show the overwhelming traffic source, and concentrations were found to be between tens and hundreds of ppb (Liu et al. 2000). In fact, driving cycle tests of in-use Chinese motor vehicles indicated that VOC emission factors were 5–10 times higher than levels in industrialised countries (Fu 2001). Interestingly, concentrations measured in Hong Kong were comparable to those in other Chinese cities (Guo et al. 2004a; So and Wang 2004). Analysis of recent measurements of VOCs within 43 cities in China (Barletta et al. 2005) results in traffic to be the biggest source. However, in some of the cities also the burning of biofuel and coal as well as natural gas leakages had a considerable influence.

2.3.4

VOCs in rural and continental background areas

Moving from the urban environment to moderately remote regions, concentrations of anthropogenic VOCs decline depending on the distance of the site to urban areas and the influence of local sources (Sexton and Westberg 1984). Concentrations are smaller than in urban areas but ratios between the species are generally in good agreement

Anthropogenic VOCs

53

4

ppb

3 C2-benzenes Toluene Benzene 2

1

0 Zurich (centre)

Zurich (suburban)

Rigi (remote)

Jungfraujoch (high altitude)

Figure 2.11 Annual mean concentrations of aromatic VOCs in 2003 at different sites in Switzerland, ranging from the polluted city centre of Zurich to the high Alpine site of Jungfraujoch.

with transported urban pollutants and their photochemical transformation (Borbon et al. 2004). As an example for the effect of dilution and degradation during transport, the development of aromatic VOC concentrations from an urban area to a background site is shown in Figure 2.11. Highest concentrations are measured in the city centre of Zurich. Values are already slightly lower at the suburban site and decline to about a 15–40% of their initial value when they reach the rural station of Rigi. Finally, at the high Alpine background site of Jungfraujoch (3 580 m asl) concentrations decrease to 5–20% of their urban concentration. Thereby, toluene and C2 -benzenes declined more rapidly than benzene because of their shorter atmospheric lifetime. However, short-term meteorological phenomena could mask this general feature. Thus, a strong increase of VOC concentrations can be found at elevated sites during summer afternoons, which is caused by convective transport of air masses from the polluted boundary layer (Helmig et al. 1996; Henne et al. 2006; Li et al. 2005; Prevot et al. 2000a). Concentrations of OVOCs in rural regions tend to grow due to photochemical production during transport and are therefore gaining in relative importance to the other VOCs. Thus, in European summer a considerable part of the measured oxygenated VOCs at rural background sites could be attributed to the secondary production from precursor VOCs (Borbon et al. 2004; Solberg et al. 1996).

2.3.5

VOCs in global background regions

Background concentrations of VOCs exhibit annual cycles, which depend on the hemispheric emissions and on the temporal behaviour of the oxidation capacity of the atmosphere. In addition, concentrations can be enhanced even at the most remote locations due to advection of polluted air masses. For practical reasons it has not been possible to continuously measure VOCs in the global background. Data coverage especially in the

54

Volatile Organic Compounds in the Atmosphere

Southern hemisphere is related to ship cruises or air plane campaigns, mostly performed in Austral summer. In view of the quantity of campaigns at background sites within the last decades it is impossible to give an all-embracing overview of the measured concentrations. However, without being comprehensive, examples of large-scale measurements or their compilation can be found in Rudolph and Johnen (1990), Koppmann et al. (1992), Singh and Zimmerman (1992), Donahue and Prinn (1993), Blake et al. (1997), Gros et al. (1998), Emmons et al. (2000), Singh et al. (2000), Lewis et al. (2001), Boudries et al. (2002) and Warneke and de Gouw (2001). VOCs in background air show an annual cycle with lower concentrations at the end of the summer because of the higher abundance of the oxidising agents (i.e. mainly OH radicals) during the warm season. The amplitude of the seasonal cycles is highest for substances which have an intermediate lifetime of days to weeks, as stable compounds do not significantly degrade during transport. Substances which have an atmospheric lifetime in the range of the transport time are partly oxygenated but still reach the remote areas in considerable amounts. On the other hand, very fast-reacting species (e.g. alkenes) are degraded almost totally during transport, which results in very low concentrations in remote areas (Hopkins et al. 2002; Lewis et al. 2001). VOC background concentrations are highest in the mid-latitude of the Northern hemisphere (30◦ –60◦ N) because of the higher amount of fossil fuel usage in the United States and Europe in comparison with other regions of the world and because the VOC lifetime is shorter than the average interhemispheric exchange time (Gupta et al. 1998). On a global level the relatively stable ethane, ethyne and propane are the most important anthropogenic NMHCs, although a considerable fraction of these substances is also emitted by biological processes (Rudolph and Johnen 1990). The abundance of these gases has been relatively stable in the last decades, as Antarctic ice core measurements showed no increase for propane and ethyne and only a minor increase (+1.6 ppt/year) for ethane since 1975 (Kaspers et al. 2004). These values are in good agreement with values of atmospheric campaigns in the Southern Hemisphere (Boissard et al. 1996; Gros et al. 1998; Singh et al. 1988) with the exception of propane being particularly high in the TROPOZ II experiment (Boissard et al. 1996). Concentrations of OVOCs normally dominate over those of NMHCs in remote oceanic air masses (Lewis et al. 2005; Singh et al. 2000), due to the oxidation of primarily emitted VOCs during transport and biological sources.

2.3.5.1

Vertical distribution

Anthropogenic VOCs are mainly emitted within the lowermost layer of the troposphere. Chemical reactions and limited transport to the higher layers of the atmosphere lead to a decline in concentrations at higher altitudes (Blake et al. 2001; Boissard et al. 1996; Rudolph 1988). The rate of decrease is higher for fast-reacting species and only relatively stable compounds can actually reach the uppermost layers of the troposphere or the stratosphere (Blake et al. 1996, 2003b; Scheeren et al. 2003; Singh et al. 1997, 2000). This general feature is, however, at times disturbed by plumes of anthropogenic emissions introduced to higher altitudes by uplifts related to meteorological processes

Anthropogenic VOCs

55

(Fischer et al. 2003; Gros et al. 2004; Jacob et al. 2003) or to topographical barriers, such as mountain chains (Henne et al. 2006; Li et al. 2005; Prevot et al. 2000b; Purvis et al. 2003).

2.3.5.2

Intercontinental transport

In the past years more and more campaigns put the spotlight onto the role of transport of air pollutants from one continent to another (Stohl and Eckhardt 2004). Thus, transport of VOCs from Asia into unpolluted Pacific air masses has been detected, leading to elevated concentrations in the higher troposphere (Blake et al. 2003a; Heald et al. 2003; Jacob et al. 2003). Furthermore, transport from the United States to Europe has been detected (Stohl et al. 2003). This issue is expected to become more and more important with emerging megacities in the tropics and Asia, with China as a special point of concern (Elliott et al. 1997).

2.4

Chemical behaviour of VOCs in the atmosphere

After their release into the atmosphere, VOCs are oxygenated by photochemical processes, which finally lead to their removal from the atmosphere. For most VOCs the process is initiated by atmospheric radicals like OH, O3 , NO3 and Cl, with the OH radical being by far the most important reactant. Only some OVOCs (e.g. formaldehyde and acetone) are susceptible to destruction by direct photolysis, as they absorb radiation in the wavelength range above 290 nm. The atmospheric lifetime of an individual VOC species is dependent on its chemical structure, the radical concentration and the intensity of solar radiation. When VOCs are degraded in polluted air masses, NO is oxygenated to NO2 , which then gets photolysed and contributes to the formation of tropospheric ozone. As this is a key issue in air pollution control, different concepts have been developed to account for the ozone formation potential of the various VOC species. Globally, the fast-reacting biogenic isoprene and terpenes are more important than anthropogenic VOCs. However, within urban environments anthropogenic VOCs are one of the dominant drivers of air pollution.

2.4.1

VOC degradation pathways

The hydroxyl radical (OH), as the most important substance involved in the degradation of VOCs, has an average global abundance of about 1.1 × 106 radicals/cm3 (Prinn et al. 2005). Its abundance is highest in summer during day hours and, conversely, lowest in winter and during the night. The importance of the Cl radical is still under scrutiny on the global scale, although indications of reactions have been found in at least one urban area (Houston) (Tanaka et al. 2003). For the alkenes also reaction with O3 becomes competitive in polluted areas, whereas the reaction of alkanes and alkenes with NO3 is mainly important during the night (Geyer et al. 2001). Reaction mechanisms and rates coefficients for the degradation of many species have been continuously reviewed by Atkinson and Arey (2003). The determination of rate coefficients for atmospheric reactions has been the result of meticulous laboratory experiments using pulsed and flow experiments. For new or exotic substances, where reaction rate coefficients are not available structure–reactivity relationship methods exist for their estimation.

56

Volatile Organic Compounds in the Atmosphere

2.4.1.1

Degradation of atmospheric alkanes

Alkanes are attacked by OH (or Cl) radicals via hydrogen abstraction, as illustrated for ethane: CH3 CH3 + OH → CH3 CH•2 + H2 O For longer alkanes, with different types of hydrogen–carbon bonds, reaction rates for the OH abstraction are decreasing with the number of hydrogen atoms attached to the same −CH2 → − −CH3 . carbon atom. Therefore, the abstraction is favoured as follows: ==CH → − The resulting alkyl radical is fast reacting with O2 to form an alkyl peroxy radical: CH3 CH•2 + O2 + M → CH3 CH2 O•2 + M In the presence of NO (>10–30 ppt) oxygen is abstracted to form an alkoxy radical: CH3 CH2 O•2 + NO → CH3 CH2 O• + NO2 The produced NO2 molecule promotes the tropospheric ozone formation, whereas the alkoxy radical, depending on the molecule, can react via the following pathways: (a) thermal decomposition, (b) isomerisation and (c) reaction with molecular oxygen (O2 ). (a)

The breaking of the carbon skeleton leads to the formation of a stable aldehyde and a smaller peroxy radical, which again can react as described above. (b) Isomerisation occurs by internal hydrogen abstraction and eventually leads to a hydroxycarbonyl molecule. (c) Reaction with O2 leads to the formation of a stable ketone via the abstraction of a hydrogen atom, which is attached to the same carbon atom as the oxygen radical. Furthermore, alkyl peroxy radicals can also attach an NO or an NO2 molecule. In the case of NO this leads to the formation of relatively stable organic nitrates, whereas with NO2 peroxynitrates are produced: CH3 CH2 O•2 + NO + M → CH3 CH2 ONO2 + M CH3 CH2 O•2 + NO2 + M → CH3 CH2 O2 NO2 + M The yields of the organic nitrates are normally higher for longer chain alkanes. Peroxynitrates (RCH2 OONO2 ) are not be mistaken for peroxyacyl nitrates RC(O)OONO2 (see below). They decompose rapidly at 298K but are more stable in colder temperature. In very clean environments the peroxy radicals can also react with other peroxy radicals (i.e. HO2 or organic peroxy radical). Reaction with HO2 results in the formation of more stable organic peroxides: CH3 CH2 O•2 + HO2 → CH3 CH2 OOH The combination with another peroxy radical is either a sink for the radicals and produces alcohols, ketones and organic acids or leads to alkoxy radicals, which reacts as described above: CH3 CH2 O•2 + CH3 CH2 O•2 → CH3 CH2 OH + CH3 CHO + O2 CH3 CH2 O•2 + CH3 CH2 O•2 → CH3 CH2 O• + CH3 CH2 O• + O2

Anthropogenic VOCs

2.4.1.2

57

Degradation of atmospheric alkenes

In the case of alkenes OH radicals preferably react by addition to the C==C double bond as illustrated for ethene: −CH•2 −CH2 − CH2 ==CH2 + OH → HO− This hydroxyalkyl radical attracts an oxygen molecule to form a hydroxyalkyl peroxy radical: −CH•2 + O2 → HO− −CH2 O•2 −CH2 − −CH2 − HO− which predominately is converted to a hydroxy alkoxy radical via its reaction with NO. −CH2 O•2 + NO → HO− −CH2 O• + NO2 −CH2 − −CH2 − → HO− The hydroxy alkoxy radical can then either react with O2 via H-abstraction to become a glycolaldehyde or by cleavage to form an aldehyde and a hydroxy alkyl radical, which can react again as described above. Furthermore, larger chained molecules have the possibility to undergo isomerisation. The second major pathway for the destruction of alkenes is their reaction via absorption of an ozone molecule (O3 ) over the C==C double bond to form an instable ozonide intermediate, which then decomposes to a carbonyl compound and a Criegee intermediate, as shown for ethene: CH2 ==CH2 + O3 → HCHO + [H2 COO]∗ Criegee intermediates can either stabilise by collision or decompose further. After stabilisation, a possible pathway leads via reaction with water vapour to the formation of organic acids. For further details see, for example, Calvert et al. (2000).

2.4.1.3

Degradation of atmospheric aromatic VOCs

The degradation of aromatic VOCs proceeds by two different pathways: (a) by abstraction of a hydrogen atom from one of the alkyl substitute groups or (b) by OH radical addition to the aromatic ring. In reaction (a) the stable product is an aromatic aldehyde (e.g. toluene leading to benzaldehyde). Reaction (b) is the predominant pathway for the degradation of aromatic VOCs. After the addition of the OH radical to the aromatic ring, molecular oxygen is added to build a cyclic hydroxy peroxy radical. In the following the ring structure is opened and formation of epoxy compounds, different saturated and unsaturated dicarbonyl radicals and finally methylglyoxal has been observed. Due to a large number of laboratory studies on the atmospheric oxidation of aromatic hydrocarbons in the past years (see the extensive review by Calvert et al. 2000), considerable improvement in the understanding of detailed chemistry has been achieved. However, ring-opening routes (b) are still speculative, whereas ring-retaining routes (a) are comparatively well understood (Hamilton et al. 2003).

2.4.1.4

Degradation of atmospheric OVOCs

OVOCs are either directly emitted by anthropogenic or biogenic sources or are produced in the degradation of VOCs, as described above. For the OVOCs the abstraction of a hydrogen atom from the carbon chain via an OH radical is the predominant reaction. Furthermore, if C==C double bonds are present also the addition of O3 is a possible initial step. The resulting

58

Volatile Organic Compounds in the Atmosphere

peroxy radicals then react as described above. For those OVOCs which have UV-absorbing groups (e.g. aldehydes, ketones, organic peroxides and organic nitrates) also the photolytic degradation can be significant. Aldehydes degrade by photolysis and by reaction with OH radicals. Whereas HCHO is degraded predominantly by photolysis, higher aldehydes react mainly with OH radicals. Products of the photolysis finally lead to the net production of HO2 , which can then be a source for OH radicals, as shown for formaldehyde (HCHO): HCHO + hν → CO + H2 HCHO + hν → HCO + H HCO + O2 → HO2 + CO H + O2 + M → HO2 + M The initial steps of the aldehyde degradation by OH radicals proceed as shown for the example of acetaldehyde: CH3 CHO + OH + O2 → CH3 C(O)O•2 The produced peroxy radical is a precursor for the PAN formation, if enough NO2 is available. The most important peroxyacylnitrate is the peroxyacetyl nitrate (PAN, CH3 C(O)O2 NO2 ), which is well known as eye irritant in emog situations. Pan and organic nitrates (see above) are relatively stable in cold temperatures and are therefore important for long-range transport as they act as reservoirs for reactive nitrogen. In the presence of sufficient NO, however, the peroxy radical of a Cn -aldehyde reacts dominantly with NO, leading to a Cn−1 -aldehyde and CO2 . For the OH radical initiated degradation of ketones the reaction proceeds by H-atom abstraction and subsequent formation of alkoxy radicals. The degradation of alcohols is mainly initiated by OH radicals. If the H-atom is abstrac−H of the CHOH or CH2 OH group, the following reaction of O2 with the ted from C− α-hydroxyradical leads to the formation of a ketone for secondary alcohols or of an aldehyde for primary alcohols. This is shown for 1-butanol and 2-butanol. OHCH2 CH2 CH2 CH3 + OH → OHCH• CH2 CH2 CH3 + O2 → CHOCH2 CH2 CH3 CH3 CH(OH)CH2 CH3 + OH → CH3 C• (OH)CH2 CH3 + O2 → CH3 C(O)CH2 CH3 Noteworthy, no NO is needed for these reactions to form stable products. −H, which leads to reactions Furthermore, OH can abstract H-atoms from other C− analogous to those for alkanes.

2.4.1.5

Lifetimes of VOCs

For the determination of the decline of VOCs in the atmosphere the following equation can be used for the description of its reaction with a radical (as shown for OH), Ct = C0 exp(−k[OH]t ) where Ct represent the concentration at the time t and C0 at the source, k is the reaction rate coefficient and [OH] is the abundance of OH radicals. k can be replaced by the pseudofirst-order rate coefficient k  = k × [OH] (i.e. reaction rate) when the [OH] concentration is assumed to be constant.

Anthropogenic VOCs

59

The atmospheric lifetime (T ) of VOCs can then be calculated by the sum of the reciprocal values of the individual rates of the radical reactions (i.e. for OH, O3 , Cl, NO3 ) and the photolysis rate constant (jA ).     + jA−1 T = kOH −1 + k −1 + k −1 + k Cl O NO−1 3

3

Thus, T denotes the time that is needed until 1/e of the initial concentration is left. Lifetimes of VOCs in the atmosphere are in the range of days to months for alkanes and alkynes, but can be as short as some hours for the most reactive alkenes, such as terpenes, isoprene and butenes (Atkinson and Arey 2003).

2.4.2

Usage of VOC reactivities for analysis of observations

The specific atmospheric reactivities of VOCs have been used to address several questions in atmospheric science. The ratio of concentrations of VOCs with different lifetimes has been used for the estimation of the air mass ‘age’. Which means by using the ratio of VOCs (with different lifetime) at the source and at the receptor together with the average abundance of the degrading radicals during transport, it is possible to deduce the travel time of the air parcel. This concept has been used in many studies, ranging from areas downwind of regional sources (Kleinman et al. 2003; McKeen et al. 1996; Roberts et al. 1984) to continental outflow (Goldan et al. 2004). However, as pointed out by McKeen and Liu (1993), this idealistic concept is influenced by non-linear processes, such as sources on the transport path and mixing of different concentrations of VOC in the background air. Vice versa this approach has been used to determine the average OH radical abundance during transport (Kleinman et al. 2003; Kramp and Volz-Thomas 1997; Williams et al. 2000; Wingenter et al. 1999). Furthermore, ratios of VOCs have been used to assess the influence of Cl radicals to the decline of VOCs in the Arctic boundary layer, as certain species have different rate coefficients for the reaction with OH or Cl (Hopkins et al. 2002; Rudolph et al. 1996). Another approach uses the fact that the spatial and temporal variability of hydrocarbon mixing ratios display an inverse power law relationship to atmospheric lifetime (i.e. variability lifetime relationship). This has been applied to calculate the average OH radical abundance over the Pacific Ocean (Ehhalt et al. 1998) and to assess the influence from sources to measured concentrations (Hopkins et al. 2005; Jobson et al. 1999).

2.4.3 2.4.3.1

Impacts of atmospheric VOCs Tropospheric ozone formation

Atmospheric VOCs influence the physical and chemical behaviour of the troposphere in several ways. Probably the most important process in relation to ‘air pollution’ is their contribution to the photochemical ozone formation. The influence of different VOC species is dependent on the spatial and temporal pattern of emissions, photochemical reaction rates and the molecule-dependent potential for producing ozone. The most widely used approaches to rank VOCs according to their ozone production ability under atmospheric

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conditions are the photochemical ozone creation potential (POCP) (Derwent and Jenkin 1991) and the maximum incremental reactivity (MIR) scale (Carter 1994). The POCP describes the amount of ozone which is built from the introduction of a VOC into the atmosphere. For the calculation of the POCP, reactions from a chemistry model are linked with trajectory models under real-world conditions. For the comparison of the ozone-building potentials of different VOCs, the POCP for a VOCi is calibrated by ratioing its absolute values with those of another VOC (e.g. ethene), which can be expressed in the following equation: POCPi = simulated extra amount of ozone due to VOCi /simulated extra amount of ozone due to ethene × 100 Applications have been the regional scale ozone formation over periods of several days under European conditions using a chemistry transport model and a trajectory model (Andersson-Skold and Holmberg 2000; Derwent et al. 1998). The MIR scale values are determined from irradiations of simplified model photochemical systems. They have, for example, been used to assess the ozone formation over periods of up to a day under optimum VOC/NOx conditions in urban environments in the United States (Carter 1994; Carter et al. 1995).

2.4.3.2

Formation of SOAs

Secondary organic aerosols (SOAs) are a major component of the total aerosol mass and can therefore contribute to the modification of the radiative balance of the atmosphere induced by particulate matters (Kanakidou et al. 2005). Apart from biogenic VOCs (e.g. terpenes) also anthropogenic aromatic VOCs have been found to contribute to the formation of SOAs (Odum et al. 1997). Thereby the atmospheric degradation of gaseous aromatic VOCs leads to the addition of a polar functional group (e.g. carboxylic acid, aldehyde, ketone and nitrate). In comparison with the initial substances oxygenated products exhibit a lower vapour pressure and higher water solubility, which favours their transfer into the particulate matter. Furthermore, oxidation products of aromatic VOCs have been found to form oligomers, which again are incorporated into the aerosol phase (Kalberer et al. 2004), leading to lower volatilities of the SOA and subsequently higher aerosol yields.

2.5

Measurement techniques

The wide range of both short- and long-lived VOC species that are of interest in atmospheric science, coupled to extremely low concentrations and a requirement often for in situ automated analysis, has led to the development of many techniques and methodologies. While some atmospheric species such as organic acids, peroxides and aldehydes are measured using high performance liquid chromatography (HPLC), the majority of organic species are analysed using capillary GC and there is an extensive literature dating back many years (Ioffe et al. 1977). In recent years, on-line MS detectors have found considerable application in atmospheric measurements of VOCs. The diversity of compounds that are of interest has resulted in almost every kind of analytical detector finding a role within VOC analysis by

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chromatography. In recent years direct introduction mass spectrometry has also been used and this in many instances has increased the time resolution with which certain species may be measured compared to chromatographic techniques. Chromatography is one of the most common methods of separating complex mixtures into its constituent components, tracing its origins back to the early twentieth century when botanist Mikhail Tswett used a liquid–solid adsorption separation to resolve many coloured components within chlorophyll. This original application gave the technique its name from the Greek for colour or ‘Chroma’ and writing ‘graphein’. From the mid-twentieth century onwards, chromatography expanded rapidly in many directions: 1940, the development of partition chromatography (for which A.J.P. Martin and R.L.M. Synge were awarded the Nobel Prize); 1952, GC, 1970, bonded particles for liquid chromatography and in 1981 the pioneering of electrically driven separations, or electrophoresis (Bartle and Myers 2002). At the heart of all chromatographic techniques, however, is a common set of principles. The mixture of components is solvated in a fluid (referred to as the mobile phase) which is used to transport the analytes through the separation system. The flow of mobile phase passes over or through an immiscible bed of material or stationary phase. Interactions occur between analytes and stationary phase, and this slows the rate at which they pass through the bed. Each analyte may interact (e.g. by adsorption) to a different extent with the stationary phase, resulting in the physical separation of the mixture into its component parts. Chromatography covers a wide range of methods and techniques so it is useful to classify these into a number of key subsets. Perhaps the most basic description that can be made is of the physical state of the mobile phase: gas, liquid or supercritical. A second classification describes the way in which the mobile phase interacts with the stationary phase. Where the stationary phase is either packed into or coated onto the walls of a length of tube it is referred to as column chromatography. The physical dimensions can vary considerably; VOC analysis can invoke columns as long as 100 m and as narrow as 50 μm whereas liquid chromatography takes place on columns of relatively modest length but up to several centimetres wide. By applying positive pressure at the start of the column, mobile phase can be driven through the column with a parabolic flow profile. Packed column chromatography involves the flow of mobile phase through a column filled with a finely powdered stationary phase, in the order of micrometres in diameter, such as a molecular sieve or silica. In capillary column chromatography, the stationary phase, usually a highly viscous liquid, is coated onto the walls of the column, and the mobile phase flows over this thin film. In GC, the mobile phase, a highly diffuse gas, is used to solvate and transport components within a mixture through or over a stationary phase, with separation occurring due to differences in the rates of migration. The mobile phase gas must have a high diffusion coefficient to ensure maximum numbers of gas–stationary phase interactions. The most common gases used in VOC GC are nitrogen, hydrogen and helium. Maximum separation efficiency is achieved using the lighter gases and as such nitrogen finds fewest applications. Highest separation efficiencies are achieved using hydrogen gas and this is over a wider linear velocity range than helium. However, the carrier gas of choice is usually helium due to the safety implications associated with hydrogen. High purity gases are used and it is normal to further purify the carrier gas with the use of a combination of oxygen, moisture and hydrocarbon traps.

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The time a compound takes to be eluted from the column is known as the retention time and is used in the identification of analytes through standard injections. In a correctly optimised chromatographic separation the analyte band will reach the detector in a narrow plug with a degree of band broadening to produce a Gaussian-shaped peak. The height or area under the peak is representative of the sample concentration giving a quantitative measurement. The overall resolving power of a chromatographic column can be given in terms of peak capacity. This is the maximum number of component peaks that can be theoretically resolved on a given column. In GC peak capacities are generally considerably larger than in liquid chromatography with typical values being 200 and 20, respectively. To completely resolve highly complex atmospheric samples would require extremely high peak capacities and selective detectors Therefore, extraction techniques and derivatisation are often used to simplify the separation required.

2.5.1

Sampling

The initial step of sample acquisition is far from being a trivial task in atmospheric measurements, and defines a major grouping of techniques into either in situ or post-acquisition analysis. The ability to store an atmospheric sample is a critical factor as to whether analysis may be performed back in the laboratory or on-site immediately following acquisition. Stable species such as methane, CO and CFCs may be successfully stored in certain sample vessels for several weeks without affecting sample integrity, and widespread measurements of these species have been performed. Each analyte however requires careful independent testing for stability, and a single storage material is not appropriate for all species – for example, CO has been shown to be problematic when stored for a longer term in stainless steel vessels. On the other hand, atmospheric organic degradation products, such as PAN (CH3 C(O)OONO2 ), are so unstable that analysis must be performed immediately. For many other important reactive species such as alkenes, monoterpenes and dimethyl sulphide (DMS), the storage of samples has been shown to lead to a degree of analyte losses due to reaction with co-sampled pollutants such as ozone and oxides of nitrogen (Helmig 1997). To overcome these problems of reaction during storage, many forms of scrubber (Helmig 1999; Sin et al. 2001; Vairavamurthy et al. 1992) have been tested to remove such species without affecting sample integrity. In other cases analytical interferences such as CO2 are removed during sampling. A great deal of VOC sampling is performed using stainless steel canisters of 1–10 l in volume (often referred to as whole air samples or ‘grab’ samples), filled either by vacuum release over a predetermined time via a flow restrictor or by pressurising the sample into the canister using stainless steel bellows or Teflon diaphragm pumps. To retain sample integrity there must be minimal interaction with the canister walls and coating methods, such as electropolishing; silica or Teflon coating are currently in use. The preparation and cleaning of canisters requires careful attention and high vacuums are often applied along with elevated temperatures and humidified zero gases for periods of hours or days. Electropolished SUMMA-canisters are often used in campaigns either in air planes or on remote locations. Applications range from very short sampling intervals, for example, in air plane campaigns (Blake et al. 1996) to long-term sampling at remote sites (e.g. EMEP, Solberg et al. 1996).

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Similar to canister samples is the use of collapsible Teflon or Tedlar sample bags. These may be filled by pump or external vacuum and returned to a laboratory for analysis using almost identical procedures to that of canister samples. While generally of lower unit cost, they often produce a greater degree of sample artefact and analyte losses. Many lower volatility species are unsuitable for collection using canister methods because of problems associated with analyte condensation onto the walls of the container. For this reason a second method of sample acquisition is widely used based on the use of a solid phase adsorbent (which is essentially a very high capacity packed column) as an analyte trap. The adsorbent used in the trap may be chosen to introduce an element of selectivity to the trapping mechanism, although in practice a trap-all approach is commonly used. Allowing air to enter an adsorbent trap purely via diffusion is referred to as passive sampling and is preferred for situations which do not require a very high temporal resolution, for example, for surveying industrial environments or workplace exposures. A huge range of adsorbent supports are commercially available (Camel and Caude 1995), ranging from high surface area (>1 000 m3 /g) carbon material (both charcoals and graphitised) with strong retention characteristics, to lower surface area (<50 m3 /g) polymerics such as Tenax TA. While being relatively low cost compared to sample canisters, care is often required in the cleaning and preparation of sample tubes. Samples may be introduced to the adsorbent tubes either dynamically over short periods of time (typically minutes) or via diffusional sampling over longer periods (typically several days). Carbon-based adsorbents are suitable for a wide range of species ranging from volatile hydrocarbons and CFCs to organic nitrates. Polymeric materials are used mainly for the concentration of lower volatility species such as aromatics and monoterpenes, although compounds as large as 2- and 3-ring polycyclic aromatic species may also be successfully trapped and thermally or solvent desorbed. Recycling of a single adsorbent allows for instrument automation, not possible with canister methods and such recycled traps now form the basis of many national monitoring programmes in the urban environment, as well as automated instruments for research in clean air environments (e.g. see http://www.epa.gov/ttn/amtic/airtoxpg.html). Sulphur species trapping is often performed via chemisorption onto gold wool traps that provide a stable matrix for the sample to be stored for reasonable periods of time (Berresheim et al. 1998). For a review of atmospheric sampling of organic compounds, see Rudolph et al. (1990b).

2.5.2

Removal of water

The inevitable presence of water in atmospheric samples and its removal prior to GC separation is a complex area. In certain circumstances its presence may be both beneficial, for example, with canister samples where it occupies the active sites on the vessel walls, and detrimental, notably through affecting either the detector or reproducibility of the separation. Alumina porous layer open tubular (PLOT) columns are particularly affected by moisture in the sample, and large changes in stationary phase affinity occur when water is introduced. Detectors such as mass spectrometers are also extremely sensitive to water introduced with the sample, and high background noise may result. Even the robust FID can be affected by injection of water, where the flame may be extinguished when water elutes from the column.

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Many forms of selective water removal exist, the simplest of which is the use of condensation traps or stripping coils. Losses of light molecular species are insignificant although condensation of higher boiling organic material may arise at very low temperatures. Inorganic adsorbents are also commonly used, notably potassium carbonate and magnesium perchlorate. Adsorbents such as these, however, have limited capacity and often require frequent regeneration or replacement. A combination of initial condensation and second stage adsorbent scrubber often provides sufficient capacity to dry a sample stream of air for many hours or days. Continuous drying may be achieved using permeation membranes such as Nafion. These forms of dryers operate by generating a steep concentration gradient across a membrane permeable only to highly polar material such as water. A counter-current of dry gas is passed around the outside of the membrane as a sheath gas and carries moisture away to waste. However, this type of drier is unsuitable for samples where the quantitation of polar materials is required.

2.5.3

Sample pre-concentration

Whether an atmospheric sample is collected using adsorbent or canister techniques, several stages of sample preparation are required prior to injection to the analytical column, since the concentration of VOCs, even in polluted environments, is too low to allow for direct loop injection on column. For volatile species collected in canister or whole air samples, analytes are removed from the canister via either internal canister pressure (for pumped in samples) or by vacuum pump (for atmospheric pressure samples) over a pre-concentration trap. A simple method of pre-concentration is to place a sampling loop either directly into or in the headspace of liquid nitrogen (Kohno and Kuwata 1991). Liquid argon has also been used since it may reduce the amount of oxygen retained in the refocusing stage. The pre-concentration zone may consist of a packed tube containing either an absorbent such as Tenax TA, glass beads or may simply be empty stainless steel tubing. Since the pre-concentration stage is at such low temperatures the majority of water vapour in the sample must be removed prior to refocusing in order to stop blockage of lines with ice. Once a sufficient volume has been collected on the refocusing trap, the trap is generally flash heated either electrically or using hot water. This results in a very sharp band of compounds being introduced to the head of the analytical column. Cryogenic methods with careful engineering can be automated successfully and used to measure at low atmospheric concentrations (Rudolph et al. 1990a). With adsorbent tube analysis the collected analytes are generally thermally desorbed either directly onto the analytical column (in the case of programmed temperature vaporisation (PTV) injection), or onto a refocusing cold trap. The desorption temperature is generally defined by the maximum temperature that the adsorbing material can support. For polymeric adsorbents this may be relatively low (<250◦ C) while carbon-based materials may support desorption at temperatures of 500◦ C or higher. If a PTV injector is used, desorption may be sufficiently rapid (>16◦ C/s) that initial phase ratio refocusing on-column is sufficient to obtain well-resolved peaks (Lewis et al. 1996). If the desorption from an adsorbent tube is relatively slow then a refocusing step is used, with a similar pre-concentration mechanism to that used with canister samples.

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Once again, water must be removed from the sample since it may affect the column or the detection system. Automated systems where a single recycled adsorbent trap is used may operate by taking air from either a local manifold or from canister collected samples. The addition of a multiposition valve upstream of the instrument can allow for fully automated canister analysis on systems designed for in situ analysis. Fully automated instruments of this kind using either cryogen or Peltier cooled adsorbent traps are commercially available (e.g. Perkin Elmer, Markes International). While the majority of species are thermally desorbed from adsorbent traps onto the analytical column a few types of compounds require solvent extraction prior to syringe injection. The analysis of some organic nitrates has been described in this way, along with higher molecular weight polycyclic aromatic compounds that may suffer from incomplete or slow release. Derivatisation can be used to selectively convert species from collected atmospheric samples to analytes more suited to GC (Cecinato et al. 2001). An example of post-acquisition derivatisation is the analysis of atmospheric alcohols, particularly of interest in Brazil where ethanol-fuelled cars are used on a large scale (Nguyen et al. 2001). Alcohols are highly polar and can stick to canisters walls or be irreversibly trapped onto absorbents. A novel technique for the analysis of alcohols involved collection of air samples in Pyrex vessels followed by subsequent reaction of alcohols with nitrogen dioxide to form alkyl nitrates. Analysis of the adducts by gas chromatography electron capture detector (GC-ECD) gives detection limits of around 1 ppb for methanol and ethanol.

2.5.4

Separation of VOCs

While a number of specific applications utilise packed columns (notably in the analysis of methane, CO and N2 O), current methods for the separation of VOC components are performed almost exclusively using capillary column GC. For a general review of the analytical science of GC, see for example Grob and Barry (2004) or chapters within general analytical science textbooks such as Skoog et al. (1998). With the introduction of Al2 O3 PLOT columns, where stationary phase is bonded to the column wall, high resolution analysis of very high volatility species is possible and many applications that historically used packed column GC are now performed using this type of hybrid capillary columns. The wide range of VOC volatilities that are encountered in the atmosphere confines each analytical system to only a limited range of species that may be completely resolved on a single column. For the most volatile NMHCs, PLOT columns are used widely for species in −C7 (Habram et al. 1998). This type of gas–solid chromatography has the carbon range C1 − a thin porous layer of a finely divided solid, usually alumina or molecular sieve, deposited on the walls of the tube. Separations on columns of this kind are via adsorption rather than phase partition, the kinetics of which are particularly rapid. As a result theoretical plate numbers in excess of 100 000 are common even for wide bore 0.53 mm internal diameter (i.d.) columns. While in principle suitable for high volatility halogenated compounds hydrogenation and dehydrogenation effects have been reported for such species on PLOT columns. The retention characteristics of PLOT-type columns are unfavourable for oxygenated compounds, where extremely strong retention, often irreversible, can occur. Due

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to the strong retention of polar species, water becomes an important interference and its presence can severely degrade the quality of PLOT separations, manifested in highly variable retention times. The very strong retention of higher boiling point species on PLOT-type columns leads to extensive peak broadening and very lengthy analysis times. Due to this, the analysis of higher molecular weight species including mono aromatics (Hamilton and Lewis 2003), VOCs, CFCs (Frank et al. 1991), HCFCs (hydrogen containing CFC replacements) and terpenoid compounds is generally performed using wall-coated open tubular (WCOT) columns (Ciccioli et al. 1992; Street et al. 1997), where the stationary phase is generally a viscous liquid such as non-polar methylpolysiloxane or slightly polar 5% phenyl-methylpolysiloxane. Typical column specifications are 0.32 mm (i.d.) × 50 m with film thickness of between 1 and 5 μm, and a temperature gradient is used to reduce band broadening in later peaks. Wide bore 0.53 mm (i.d.) columns are also used where direct thermal desorption from a pre-concentration trap to the analytical column is applied. The rate of generation of theoretical plates (and hence peak capacity) on columns of this type are lower than for PLOT types, and as a result to obtain full resolution of some species (e.g. HCFC mixtures in the atmosphere), columns as long as 100 m have been reported. This technique is the most −C12 ) and semi-volatile organic comwidespread method of analysis of mid-range (∼C4 − pounds in the atmosphere. Columns are highly durable and can be used in both field and laboratory measurements. To improve the retention and separation of some volatile VOCs (those that fall between standard wall-coated siloxane columns and PLOT columns) without use of sub-ambient cooling, phase thicknesses up to 15 μm have been reported. Band broadening effects through stationary phase diffusion become significant with films of this thickness and this approach has not been widely adopted. Gas phase species with high molecular weights such as naphthalene, fluorene and anthracene may be separated efficiently on non-polar columns with film thicknesses of typically 0.25–0.5 μm. Since MS is commonly the detector, very low bleed columns are desirable, although this is difficult to achieve, given the film thicknesses typically employed. New stationary phases have been developed to extend the range of compounds amenable to capillary GC analysis. Adding polar functionalities to the methylpolysiloxane backbone can create stationary phases of different polarities and selectivities. For example, 14% cyanopropylphenyl 86% methylpolysiloxane columns are moderately polar and have been used in the analysis of hydrocarbons, oxygenates and CFCs. In addition, a number of speciality phases are available that allow the separation of oxygenated material by capillary GC. Since such species are generally at low concentration in the atmosphere, they often suffer from co-elution with more abundant primary emitted VOCs. While derivatisation followed by HPLC is still a most common method of carbonyl analysis, the use of mixed phase porous layer capillary columns is an emerging technique. An example is the Varian CP-LOWOX column, which uses a multi-layer column coating process to produce an extremely polar column with high stability. Sample pre-concentration can be via standard carbon adsorbent methods and since no derivatisation stages are required, minimum detectable amounts are greatly improved and sampling volumes drastically reduced. Highly polar wall-coated stationary phases have also been developed using polyethylene glycol (WAX) and can be used to separate oxygenated analytes such as alcohols, free acids and esters.

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GC separations are in general terms performed over timescales of perhaps one sample per hour, but that is not necessarily the limit of performance that can be achieved. Using narrow bore columns and fast temperature ramps (i.e. 40◦ C/min), commonly available on modern GC instrumentation, it is possible to perform much faster GC separations. If sampling, which is often a rate-determining step in the analytical procedure, can be undertaken in parallel, there are no great barriers in reducing cycle time to perhaps 5 or 10 min. This has been used to develop GC-based instruments which can be used from research aircraft (Singh et al. 2004; Whalley et al. 2004).

2.5.5

Detection

As highlighted in the previous section, even the highest resolution capillary column often has insufficient peak capacity to resolve all VOCs in a typical atmospheric sample. Since the introduction of analyte selectivity in the trapping and preparation stages is not always possible, selectivity in detection is a very useful tool for simplifying atmospheric samples. However, it is also often convenient to use a simple detector that gives a universal or groupspecific response. For an overview of possible detectors for GC, see Hill and McMinn (1992). The FID is in general terms by far the most commonly used detector in GC, since it offers high sensitivity, extremely wide linearity and very good long-term reliability and response. Using well-cleaned fuel gases coupled to low noise electrometer circuitry, it is possible to determine amounts down to as low 1 pg/s of eluting peak. With a typical sample volume of 1 l, detection limits for individual species may therefore be in the low ppt range (10−12 ). Calibration can be performed with relative ease (and in some cases calculated from standard response characteristics) but the complexity of samples can make peak identification difficult when co-elutions occur. To overcome this lack of selectivity, analytical methods for alkene and aromatic analysis using a selective response from photoionisation detection (PID) and the reduction gas detector (RGD) have been proposed, although they are not as widespread as FID techniques. For detectors such as the PID, it is the need to regularly calibrate for decreasing bulb/ionisation intensity which accounts for its limited usage. The ECD was specifically designed for organohalogen measurements in the atmosphere and appears in many published methods in the literature. It offers high sensitivity to electrophilic compounds with almost no response for hydrocarbon species. GC–ECD measurements require careful calibration due to the great variation in response to individual halogenated species, although their high stability allows gas standards to be used over many years. Some halogen-containing species of atmospheric interest (e.g. CH3 Cl, CHF2 Cl, CH2 Cl2 ) have a relatively poor ECD response and the use of the oxygen-doped ECD to enhance their response has been successful. Although MS detection is now the most widespread detection method for these compounds, the ECD still finds usage (Rivett et al. 2003; Urhahn and Ballschmiter 2000). Selective detection of oxygenated VOCs is difficult, due to low response in both ECD and FID detectors, and low molecular weight fragmentation in MS. Elemental specific detection such as AED offers some potential in oxygenate analysis although sensitivity is poor at around 100 pg/s. Detectors such as the helium ionisation detector (HID), produce a non-selective, high sensitivity response to these types of compounds, but require full

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chromatographic resolution since they also respond to every other type of VOC (Hunter et al. 1999). Mass spectrometry offers obvious additional analytical resolution, and bench-top GC–MS technology is now at a very advanced stage in terms of reliability, self-calibration, tuning and automation. The eluent from the GC column is transferred to the ion source directly via a gas tight heated transfer line making this a simple and convenient hyphenated technique, without the mobile phase evaporation problems associated with LC-MS. Ions can be produced using a number of different ionisation methods, the most popular being electron (EI) and chemical ionisation (CI). For an overview of MS techniques commonly found in use with atmospheric applications, see for example Hoffmann and Stroobant (2002), although very many other texts are available. In EI, an electron beam, operated at about 70 eV, is fired at molecules as they enter the ion source. The high-energy beam transfers large amounts of energy to molecules, which can break up into smaller ions, releasing excess energy. Fragmentation patterns can give important information about the structure of a molecule, an example being a series of m/z ions that are spaced 14 amu apart, indicating the presence of a hydrocarbon chain. However, the spectral information obtained from GC–MS of many VOC species (in particular hydrocarbons-based compounds) often leads to highly similar fragmentation patterns and assists little in the identification of isomeric species. Similarly, identification of monoterpene species can only be confirmed through a combination of both spectral information and retention time data. It is often advantageous to obtain the molecular ion to facilitate identification. In some cases, EI completely fragments the molecule and the molecular ion is absent from the chromatogram. CI produces ions using an ionised reagent gas (commonly methane, isobutane or ammonia) to transfer protons to (or from) the analyte resulting in the formation of a pseudo-molecular ion. In the positive ion mode the reagent gas donates a proton to the analyte to form [M + H]+ and in the negative mode it abstracts a proton to form [M − H]− . This is a softer form of ionisation and produces less fragmentation in the ion source. Ions are accelerated by the accelerating plate into the mass analyser, which separates ions according to their mass-to-charge ratio (m/z). The quadrupole mass spectrometer has now become one of the most widely used mass spectrometers because of its ease of use, small size and relatively low cost. Mass separation in a quadrupole mass filter is based on achieving a stable trajectory for ions of specific m/z values in a hyperbolic electrostatic field. Operating currently available quadrupole mass spectrometers in full scan mode has only moderate sensitivity for hydrocarbon-like species perhaps to the 30–50 ppt level. Where GC–MS is particularly strong is in the measurement of halogenated species in the atmosphere (Oram et al. 1996). While hydrocarbon fragmentation is often very similar with little abundant parent ion, many halogenated compounds give highly unique MS fragmentation with abundant large m/z fragments. Long-term measurements of species such as CFCs and their replacements has been performed by GC–MS instruments for many years within the world-wide AGAGE network (Prinn et al. 2000).

2.5.6

Calibration techniques and quality assurance

The quality of data required in any analytical system is of great importance, particularly if any conclusions are to be drawn from the data, such as air quality statistics and model

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simulations. In GC systems it is normal to use a number of quality control methods, but the most important within this set is the use of calibration standards. A Retention Time Standard, which is a qualitative mix of known components (such as hydrocarbons), is used initially to determine the chromatographic peak retention time and if MS is present, the instrument specific mass spectra. It can be used during instrument optimisation to ensure maximum peak separation. In addition, it can also be used during routine analysis to ensure there is no retention time shift, which could be due to factors such as changes in water content and flow blockages. A primary calibration standard is used to determine the response of the GC system to specific analytes and therefore allows quantification. This primary calibration standard is generally a gas standard, made up in an inert matrix such as zero air, or nitrogen, containing the compounds of interest at known and typical concentrations. Standard Reference Materials (SRMs) and Certified Reference Materials (CRMs) are available for some VOCs from speciality gas suppliers and are referenced to a national standard such as the National Institute of Standards and Technology (NIST) in the United States or the National Physical Laboratory (NPL) in the United Kingdom. Compressed gas reference materials such as these are generally prepared via a multiple gas dilutions of a high concentration mother standard, whose VOC content is known accurately from gravimetrical measurement. At each dilution step accuracy may be degraded and there is a tendency for such synthetic standards to be relatively high compared to ambient VOC concentrations. Even though calibration standards with concentrations approaching those of real air samples are preferable, because they reduce any errors associated with dilution during sampling of the standard or non-linearities in instrument response. Gas standards should also ideally be humidified to ambient levels to ensure there are no water-dependent effects on sampling and response. Given the requirements highlighted above, an ideal working standard is the use of ambient compressed air itself. This works well for stable species such as halocarbons (Prinn et al. 2000), but is more problematic for reactive hydrocarbons. The frequency of calibration is dependent on application and detector. An FID is considerably more stable over time than a quadrupole mass spectrometer, the former requiring a check on calibration perhaps every couple of days, the latter every couple of samples. The intercomparison of GC instruments around the world is also an important aspect of quality control. The AGAGE instruments are calibrated on-site using alternate analysis of ambient air and a 35 l standard tank, with standard tanks lasting up to ten months and then calibrated relative to a central laboratory (currently the Scripps Institution of Oceanography (SIO) California, USA). For a review of calibration methods used at GAW sites, see Plass-Dulmer et al. (2002). A number of international intercomparison exercises have been carried out for the observations of VOCs. The Accurate Measurement of Hydrocarbons in the Atmosphere −C9 (AMOHA) project was developed to evaluate current GC methods of analysis of C2 − NMHCs across Europe (Slemr et al. 2002b). Calibration gas standards were prepared by the UK NPL for all laboratories, to reduce variability associated with the standard, and analysis was carried out in 12–14 laboratories over 4 years. Agreement was generally good for high concentration standard mixtures but disagreement increased as lower concentration, real air samples were analysed. However, some improvements were achieved over the course of the project. Broadly similar finding arose from the U.S. led NOMHICE intercomparison project (Apel et al. 2003). A number of intercomparisons between different

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analytical measurements including GC-FID/ECD/MS, proton transfer reaction mass spectrometry (PTR-MS) and open path FTIR (OP-FTIR) have been made (Christian et al. 2004). Good agreement between GC and OP-FTIR and between GC and PTR-MS was achieved for a number of compounds. However, differences were found for sticky compounds (e.g. formic acid) and compounds with a low proton affinity (e.g. formaldehyde) with OP-FTIR and PTR-MS having considerable advantages. It is important for real samples to take into account the great difference in time resolution between these methods. In the future, however, it seems certain that validation of techniques via intercomparisons of standards will increase.

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Volatile Organic Compounds in the Atmosphere Edited by Ralf Koppmann Copyright © 2007 by Blackwell Publishing Ltd

Chapter 3

Biogenic VOCs Allison H. Steiner and Allen L. Goldstein

3.1

Introduction

Biogenic volatile organic compounds (VOCs) are carbon-containing compounds emitted naturally from the Earth’s surface to the atmosphere. This general categorisation includes a wide range of organic species emitted from vegetation, soils and the oceans, and generally excludes methane, carbon monoxide and carbon dioxide (Kesselmeier and Staudt 1999). This chapter reviews current knowledge on the atmospheric budget of these compounds, from the origin of emission to their fate in the atmosphere. The first biogenic VOC emissions specifically identified to be important for atmospheric composition were isoprene (Sanadze 1957) and terpenoid compounds (Went 1960). Subsequent studies indicated that a wide variety of organic species are routinely emitted from vegetation (Zimmerman 1979). The discovery that isoprene can be the dominant VOC impacting regional tropospheric photochemistry and ozone production (Chameides et al. 1988; Trainer et al. 1987) dramatically enhanced the amount of research focusing on biogenic VOCs. As research has progressed over the last few decades, the understanding of the types and quantities of these emissions has grown significantly. Multiple review articles have been published that can provide a scientific history of the topic (e.g. Fehsenfeld et al. 1992; Fuentes et al. 2000; Kesselmeier and Staudt 1999; Wiedinmyer et al. 2004). The importance of biogenic VOCs is twofold. First, ecophysiologists investigate biogenic VOCs to elucidate the mechanism of plant photosynthesis and metabolism. Secondly, they are the dominant reactive organic compounds in the troposphere, making them important to atmospheric chemists and climatologists. Many species of biogenic VOCs are rapidly oxidised in the atmosphere, playing an important role in atmospheric chemistry and the formation of ground-level ozone. Some oxygenated biogenic VOCs are photolysed in the upper troposphere, providing a source of HOx radicals and impacting the oxidation capacity of the atmosphere at large scales. Additionally, the oxidation of many biogenic VOCs can yield products that undergo gas-to-particle conversion, leading to the growth of secondary organic aerosols that affect the Earth’s energy balance and thus surface climate. The intersection of these two fields has resulted in a wide body of research on scales ranging from biological synthesis at the molecular level to atmospheric chemistry at the global level. In this chapter, we will review and discuss the state-of-the-science of biogenic VOC research. We define the chemical speciation of identified biogenic VOCs and present their known sources from the molecular level up to the landscape level (Section 3.2), and then present current estimates on the magnitude of these emissions and how they are modelled at the leaf, regional and global scales (Section 3.3). Next, we discuss their atmospheric distributions

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(Section 3.4) and their impact on atmospheric chemistry (Section 3.5). Finally, we discuss mixing ratio and flux measurement techniques used to constrain the atmospheric budget (Section 3.6).

3.2 3.2.1

Sources of biogenic VOCs Chemical speciation

Global emissions inventories as recent as the 1990s categorised biogenic VOC emissions into isoprene, monoterpenes and a general grouping of ‘other VOCs’ (Guenther et al. 1995). In early inventories, where little information was available on the chemical species emitted or the specific sources of the emissions, this term distinguished between ‘other reactive’ VOCs and ‘other VOCs’ with lifetimes greater than or less than 1 day, respectively (Guenther et al. 1995). Here, we re-categorise biogenic VOCs to reflect the most recent research by adding two categories: oxygenated VOCs (oxVOCs) and very reactive biogenic VOCs (VR-BVOC) (Holzinger et al. 2004). Table 3.1 shows examples of the chemical formulas and structures of these categories, and they are described in further detail below. The most studied group of biogenic VOCs is the isoprenoid compounds. Isoprenoids are composed of C5 (or hemiterpene) building block structures. The most well-characterised compound of this category in terms of biogenic emissions and atmospheric chemistry is isoprene (2-methyl-1,3-butadiene; C5 H8 with structure shown in Table 3.1), due to its copious emissions, high reactivity and relatively early discovery. Sanadze (1957) and Rasmussen and Went (1965) made the first measurements of isoprene emitted from foliage and found that it was released from several types of vegetation species. Since this discovery, isoprene emissions have been measured at the leaf and canopy levels in hundreds of ecosystems. Terpenoids are a large class of compounds containing combinations of the C5 isoprenoid structure. These include C10 compounds known as monoterpenes, C15 compounds or sesquiterpenes and the larger C20 , C25 and C30 molecules. Up to 5 000 different terpenoid structures have been identified in emissions from vegetation, of which 14 common species have been determined as dominant in biogenic VOC emissions (Geron et al. 2000a). The most abundant of these include the monoterpenes listed and illustrated in Table 3.1. Terpenoid emissions were first noted by Went in 1960, with a prescient hypothesis that these biogenic emissions could lead to the haze often seen in rural areas such as the Great Smoky Mountains. Because of their wide range of reactivities and their differences in propensity to form secondary organic aerosols, recent studies have underscored the importance of speciating monoterpenes for understanding their impacts on atmospheric chemistry and climate (Hallquist et al. 1999). Regarding biogenic emissions and their importance in atmospheric chemistry, monoterpenes are the most studied of the terpenoid category. They have a C10 structure that enables the molecule to be quickly oxidised in the atmosphere, yielding lifetimes ranging from several minutes up to a day. Monoterpenes have double bonds that can be contained inside a ring structure (endocyclic) or outside the ring structure (exocyclic). Some monoterpenes are a combination of exocyclic and endocyclic bonds, or they can be open-ended structures. Most importantly, the position of the double bond can affect the reactivity of the molecule, which can greatly influence atmospheric chemistry. Sesquiterpenes (C15 compounds) have

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Table 3.1 Categories of biogenic volatile organic compounds Category

Chemical formula/definition

Example compound/structure

Other compounds

Isoprenoids

C5 compounds

Isoprene

Methylbutenol

Exocyclic

C10 H16 Double bonds outside ring structure

β-Pinene

Camphene, sabinene, p-cymene

Endocyclic

C10 H16 Double bonds inside ring structure

α-Pinene

3 -Carene, α-terpinene, γ -terpinene

Combination

C10 H16 Double bonds inside and outside ring structure

Limonene

Terpinolene, β-phellandrene

Open-ended

C10 H16 No ring structure present

Myrcene

Ocimene

Sesquiterpenes

C15 H24

β-Caryophyllene

α-Humulene

Terpenoids

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Table 3.1 (Continued) Category

Chemical formula/definition

Oxygenated

C10 compounds containing oxygen

Methyl chavicol

C and O containing compounds

Acetone

oxVOCs

Example compound/structure

Terpenoid or oxVOCs with lifetimes on the order of minutes to hours

α-Thujene

O

Acetaldehyde, methanol, ethanol, methylbutenol

O

H 3C VR-BVOCs

Other compounds

β-Caryophyllene

CH3 β-Farnesene, α-terpinene

a more elaborate ring structure than monoterpenes and often contain several double bonds, leading to rapid oxidation after they are emitted to the atmosphere. Additionally, some terpenoid compounds can be oxygenated, such as methyl chavicol. Table 3.1 shows some examples of structures and compound names for these different types of terpenoid compounds. Oxygenated biogenic VOCs, or oxVOCs, have received more attention in the past decade as technology to measure their emissions and atmospheric mixing ratios has improved. oxVOCs are defined as carbon-based compounds containing an oxygen atom, including alcohols, ketones, esters and ethers. While oxVOCs have been known for decades to be emitted from plants (Zimmerman 1979), the biogenic contribution to their atmospheric budgets have been investigated only recently (Goldstein and Schade 2000; Heikes et al. 2002; Jacob et al. 2002; Schade and Goldstein 2001; Singh et al. 1994, 2000). Dominant observed oxVOC species include methanol, ethanol, methylbutenol (MBO), formaldehyde, acetaldehyde and acetone. Biogenic emissions are the dominant source of oxygenated

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compounds in the atmosphere, having a significant impact on atmospheric chemistry. The importance and emission of oxVOCs are thoroughly reviewed in Fall (2003) and discussed at length in Chapter 4 of this book. In this chapter, the focus is on biogenic oxVOC emissions of MBO, methanol, ethanol, acetone and acetaldehyde, as these are the most studied of the oxVOC biogenic species. In addition to the terpenoid and oxVOC compounds discussed above, a wide range of gases can be emitted from vegetation, including alkanes, alkenes, organic acids, carbonyls, esters and ethers (Kesselmeier and Staudt 1999). Past literature has grouped these compounds together with oxygenated compounds into the ‘other VOC’ category, but recent research provides further speciation and quantification of these important emissions. Analyses from several field measurement campaigns have recently suggested that a large number of terpenoid compounds are reacting quickly within the forest canopy, leaving them unaccounted for in fluxes measured escaping from the canopy. Ciccioli et al. (1999) discussed the importance of these compounds, particularly the large emissions of β-caryophyllene in a blooming orange grove in Spain. In this study, some monoterpenes were observed in leaf emissions, but could not be detected above the canopy. Holzinger et al. (2004) measured oxidation products in and above a ponderosa pine ecosystem and found evidence of very reactive terpenoid compounds being oxidised before escaping the forest canopy, and termed these emissions as VR-BVOCs. VR-BVOC describes difficult-tomeasure, short-lived compounds that are likely to be oxidised before escaping the forest canopy. A listing of a few of the possible VR-BVOCs is included in Goldstein et al. (2004) and Table 3.1, and includes compounds such as α- and β-farnesene, α- and β-terpinene, α-humulene, β-carophyllene and terpinolene. In addition to direct measurements, there are also several studies that provide indirect evidence of the importance of VR-BVOCs. Kurpius and Goldstein (2003) found that the main process responsible for ozone deposition to a ponderosa pine forest was gasphase chemical reaction, and that this loss scaled with temperature in the same manner as monoterpene emissions. Furthermore, Goldstein et al. (2004) found that the chemical loss term increased significantly during forest thinning, a mechanical disturbance that increased monoterpene emissions in the same proportion as ozone loss. In a different ecosystem, Di Carlo et al. (2004) found that the measured biogenic VOCs could not account for all of the OH reactivity in a deciduous forest. They suggest that unmeasured biogenic VOCs were responsible for the missing reactivity because it scaled with temperature in the same relationship as biogenic VOC emissions. These VR-BVOC emissions are important because they act as sources and sinks for atmospheric radicals and provide a source of oxidation products that may persist in the gas phase or partition into secondary organic aerosols. Here, we define VR-BVOC as compounds with atmospheric lifetimes of minutes or less.

3.2.2 3.2.2.1

Mechanisms for biogenic VOC emissions Isoprene formation

In the past decade, a significant amount of progress has been made in understanding the isoprene synthesis process. Two recent papers have thoroughly reviewed the topic of isoprene formation (Sanadze 2004; Sharkey and Yeh 2001), and here we briefly summarise current knowledge of the formation mechanisms. Isoprene is produced via de novo synthesis

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Cytoplasm MVA pathway 3-Acetyl-CoA

Chloroplast MEP pathway

HMG-CoA

GA-3P + pyruvate Mevalonate 1-Deoxy-D-xyulose-5-P DMAPP +IPP

IPP

IPP

DMAPP

IPP isoprene

+IPP

+IPP FPP

?

DMAPP

Sesquiterpenes Monoterpenes

Figure 3.1 Isoprene formation pathway, based on Lichtenthaler 1999 and Karl et al. 2002.

(Delwiche and Sharkey 1993; Sanadze et al. 1972), meaning that isoprene emissions occur almost immediately after formation. Initially, the isoprene formation pathway was thought to follow the mevalonic (MVA) pathway, which produces isoprene through acetyl-CoA. While this pathway is responsible for the synthesis of isoprenoids in the cytoplasmic space, recent research shows that a different pathway produces isoprene in the plant plastids (Lichtenthaler et al. 1997; Zeidler et al. 1997). This plastidic pathway is known as the methylerythritol phosphate (MEP) pathway (reviewed in detail in Lichtenthaler, 1999, and shown in Figure 3.1), and produces the main isoprene precursor, dimethylallyl pyrophosphate (DMAPP). DMAPP is then catalysed by isoprene synthase to form isoprene (Silver and Fall 1991, 1995). Even with these recent advances in understanding the formation pathway of isoprene, the biological regulatory controls are still uncertain. It is believed that production is likely regulated by some combination of isoprene synthase activity and DMAPP substrate production (Rosenstiel et al. 2002). Several investigations have attempted to determine the sources of carbon utilised for isoprene formation. Part of the motivation for understanding the synthesis mechanisms and sources of carbon is to illuminate the biological function of isoprene. Recent experiments have used isotopically labelled carbon dioxide to determine the percentage of isoprene’s carbon derived from the photosynthetic cycle. Based on these studies, it is now well established that chloroplastic carbon accounts for about ∼80% of the isoprene carbon under normal conditions with a lesser percentage under stress conditions (Affek and Yakir 2003; Karl et al. 2002; Kreuzwieser et al. 2002; Loreto et al. 2004; Schnitzler et al. 2004). Isoprene synthesis is therefore closely linked to the photosynthetic cycle, but other carbon reserves can be tapped to produce isoprene even when the photosynthetic cycle is operating under drought or temperature stress.

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There is still a wide range of speculation about the biological purpose of isoprene emissions, which often account for a few percent of the carbon photosynthesised by the plant. Possible roles of isoprene include providing thermotolerance for vegetation, acting as an antioxidant, or providing a mechanism to release excess energy and/or carbon. Proponents of the thermotolerance hypothesis suggest that the emission of isoprene increases the stability of protein membranes in the chloroplasts of leaves (Sharkey and Singsaas 1995; Singsaas et al. 1997), although subsequent studies could not reproduce this thermotolerance with statistically significant results (Logan and Monson 1999). Others have hypothesised that isoprene can quench ozone to below the toxic level for plants (Loreto et al. 2001b) or protect against damage by singlet oxygen radicals (Affek and Yakir 2002). A third hypothesis is that isoprene could be a mechanism to shunt excess energy (as ATP) or carbon from the leaves (Logan et al. 2000). A new hypothesis by Rosenstiel et al. (2004) suggests that isoprene acts as a metabolic ‘safety valve’ by keeping the carbon substrate and intermediate species (such as DMAPP) at optimal levels. Because of these conflicting theories, the biological purpose of isoprene production in plants remains an active area of research.

3.2.2.2

Terpenoid formation

Monoterpenes are formed from the reaction of DMAPP and IPP (see Figure 3.1), and their formation has been likened to the merging of two five-carbon branched chains (Fuentes et al. 2000). There are two pathways of monoterpene formation that depend on the nature of the emissions. Temperature-dependent terpenes are usually synthesised in leucoplasts or the cytosol by the mevalonate pathway and then stored in secretory organs in the plant such as glandular trichomes and resin ducts (Kesselmeier and Staudt 1999; Loreto et al. 2001a; Niinemets and Reichstein 2002). The terpene formation mechanism was first speculated upon by Yokouchi and Ambe (1984) and later confirmed with observations (Kesselmeier et al. 1997; Staudt and Seufert 1995; Steinbrecher 1989). The monoterpene storage structures can be internal to the leaf (such as a resin duct in a pine tree) or external (such as a glandular trichrome on the surface of mint leaves). It is believed that these compounds play a role as a defence mechanism of the plant (hence their storage in specialised structures), and specific monoterpenes are known to both attract and repel insects. A second formation pathway occurs if the monoterpene emissions are both light and temperature dependent. Like isoprene formation, this production is linked to the photosynthetic cycle and forms monoterpenes in the chloroplasts via the MEP pathway (Loreto et al. 2001a). These light- and temperature-dependent emissions likely perform a similar biological function as isoprene. Higher-order terpenoid emissions, such as sesquiterpenes and diterpenes, are formed from the similar C5 units as isoprenoids (IPP and DMAPP), as shown in Figure 3.1. At this time, it is believed that sesquiterpenes are formed via the mevalonate pathway in the cytosol (Eisenreich et al. 2001). Different terpenoid emissions are likely to have varying functions at varying scales, including protection against cellular damage at the tissue level, acting as antioxidants at the leaf surface level and acting as signal compound for insects and animals at the ecosystem level (Holopainen 2004).

3.2.2.3

oxVOC formation

Here we provide a brief review of the mechanisms of biogenic oxVOC formation; these pathways are discussed in great detail in Fall (2003). In particular, we focus on methanol,

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ethanol, MBO, acetaldehyde and acetone as the most information is available about the formation and emission of these compounds. Methanol emissions were first noted by MacDonald and Fall (1993) and can be emitted in fairly large amounts from terrestrial vegetation. One known source of methanol is from the reactions of pectin in plant cell walls (Galbally and Kirstine 2002), but other sources from terrestrial vegetation are likely as well (Heikes et al. 2002). Emissions of methanol can also occur due to plant wounding (Warneke et al. 2002) or from decomposing leaf and soil litter (Baker et al. 2001; Schade and Goldstein 2001). Ethanol, another potentially important emission, can be produced in the plant under two possible scenarios (Schade and Goldstein 2002). The first is the ‘flooding scenario’, where ethanol is produced via fermentation in submerged roots and then transported to other portions of the plant, later to be emitted via leaf tissue (Kreuzwieser et al. 1999). The other possible mechanism for ethanol emissions is the production in the leaf from fermentation under stress conditions, such as drought or damaging trace gases such as ozone (Schade and Goldstein 2002). MBO is formed in a similar mechanism to that of isoprene, despite the fact that they are emitted from different types of vegetation. As with isoprene, MBO is formed via the MEP pathway with DMAPP as the precursor (Fall 2003). Once DMAPP is formed, a MBO synthase enzyme converts DMAPP to MBO (Fisher et al. 2000). The biological functions of MBO are also unclear, though are expected to play a similar role as isoprene, given their similarities in chemical structure. Acetaldehyde is often emitted during light-to-dark transitions in plants (Holzinger et al. 2000; Karl et al. 2002), and also as a by-product of the ‘flooding scenario’ described above (Kreuzwieser et al. 1999). During light-to-dark transitions, pyruvic acid rises in the leaf and is then catalysed by a safety value to convert excess pyruvate to acetaldehyde, which is often ‘leaked’ from intercellular space. The other possible mechanism for acetaldehyde formation is by the oxidation of root-produced ethanol, which occurs in the leaves and can emit acetaldehyde under certain instances (Kreuzwieser et al. 1999). Acetone, also emitted in large quantities from vegetation, can form by several enzymatic pathways within the leaf. It can be emitted during leaf wounding or from light-dependent or light-independent responses in the leaf (see Fall 2003 for a detailed review of these pathways).

3.2.2.4

Mechanisms of VR-BVOC formation

As the category of VR-BVOCs and specific compounds therein has been recently delineated, there is little evidence for formation or biological function. Here we categorise VR-BVOCs primarily to include terpenoid compounds based on several measurement studies (Ciccioli et al. 1999; Goldstein et al. 2004; Holzinger et al. 2004). Most C10 terpenes are formed from DMAPP and IPP, and sesquiterpenes are formed from DMAPP and two IPP molecules. Different types of terpenes and sesquiterpenes require different enzymes to form the end products; therefore, it is believed that specific enzymes are required for different terpenoid formations (Iijima et al. 2004). We would expect the other reactive terpenoid and sesquiterpene compounds to play similar ecological roles as the monoterpene emissions. The presence of these compounds in plants has long been studied by wood products chemists (Zavarin 1968), but investigations on their release into and impact on the atmosphere are preliminary.

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3.2.3

Release of volatiles

Once an organism has overcome its biochemical constraints on production, biogenic VOCs must overcome the physiological constraints and travel from the intercellular air spaces and storage structures into the atmosphere. There are three possible foliar release mechanisms: 1. exit of gases via the stomates; 2. non-stomatal exit via diffusion through the leaf cuticle or woody portions of the plant; 3. release of compounds from the storage pools in plants.

3.2.3.1

Stomatal exit

Compounds that are not stored in specialised structures in the leaf can exit from the intercellular air space to the atmosphere through the stomata. In theory, this exit should be controlled by changes in the stomatal aperture (conductance); however, early studies showed that isoprene emission is not limited by the opening and closing of the stomates (Fall and Monson 1992; Monson 1989; Tingey et al. 1981). This led to speculation that isoprene was unaffected by stomatal conductance because of the large gradient between the intercellular and atmospheric mixing ratios. Fall and Monson (1992) postulated that this gradient was so great that isoprene could overcome the diffusive limitations incurred by the nearly closed stomata. The stomatal flux of gases from the leaf can be described by the following relationship: F = Gs P

(3.1)

where F is the emission flux of the biogenic VOC compound, Gs is the stomatal conductance of that compound, and P is the partial pressure gradient between the atmosphere and the inside of the leaf. In the past, it was assumed that if Gs decreased, the increase in P could overcome any stomatal control restricting release of VOCs. Other compounds, such as α-pinene and light-dependent monoterpenes, have been observed to exhibit a similar lack of stomatal control (Niinemets and Reichstein 2002). However, stomatal closure can reduce the emission of other oxVOCs, including methanol (Nemecek-Marshall et al. 1995), ethanol (Schade and Goldstein 2002), carboxylic acid (Kesselmeier et al. 1997) and other organic acids (Gabriel et al. 1999), yet the mechanistic reason for this control was not understood in these observational studies. Niinemets and Reichstein (2003a, 2003b) hypothesised that varying stomatal controls on biogenic VOC emissions are related to the partitioning between gas and liquid phases in the intercellular air space. If compounds are highly volatile with a high Henry’s law constant (H), then gas-phase mixing ratios build up immediately in the leaf and overcome stomatal resistance. However, if compounds have a low H and are likely to reside in the liquid phase, emissions will not occur until a significant gas-phase mixing ratio has accumulated. Niinemets and Reichstein (2003b) use this hypothesis to explain bursts of emissions shortly after stomatal openings.

3.2.3.2

Non-stomatal exit

Early VOC studies investigated the possibility of emissions through the cuticle of the leaf (Fall and Monson 1992; Guenther et al. 1991), though more recent evidence indicates that

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this may have resulted from leaf wounding during the experiments (Loreto et al. 1996b). In general, however, non-stomatal emissions are considered to be quite small in comparison with emissions through the stomata. Air diffusion rates through the cuticle are a factor of eight times less than diffusion through the stomata and about four times less for water (Niinemets and Reichstein 2003a). Therefore, most work assumes that the emissions are emitted via the stomata but the possibility remains that small amounts could be emitted through the cuticle.

3.2.3.3

Exit via mechanical disturbance

Another common exit pathway for biogenic VOCs is via mechanical wounding. Emissions have been measured as a result of mechanical disturbance and can last long after the physical disturbance has occurred (Rasmussen and Went 1965; Tingey et al. 1991). This can occur with natural perturbations such as insect damage or herbivore feeding (Litvak et al. 1999) or from anthropogenic influences such as pasture cutting (Kirstine et al. 1998) and forest thinning (Goldstein et al. 2004; Schade and Goldstein 2003). These emissions can include an increase in existing emissions (as in the case of monoterpene emissions increasing with herbivory feeding or forest thinning), or create emissions of new biogenic VOCs, such as the release of hexanal and hexane family VOCs after plant wounding (Fall et al. 1999). When leaves are broken or disturbed, ducts or trichromes that store biogenic VOCs can be broken open and the compounds are volatilised upon contact with air (Litvak and Monson 1998, Loreto et al. 2000). They can also be volatilised directly from the storage structures, depending on temperatures and the monoterpene pool mixing ratio (Lerdau et al. 1997). Often, this type of release can cause experimental artefacts with leaf-, branchor plant-level sampling methods, where the process of enclosing vegetation for measurements can damage portions of the plant tissue and cause emissions that might not occur otherwise.

3.2.4

Emission sources of biogenic VOCs

The predominant source of biogenic VOC emissions is from the foliage of terrestrial vegetation. This includes natural vegetation such as trees, shrubs, grasses, ferns and mosses, as well as anthropogenically induced vegetation such as crops and urban landscapes. Other minor sources such as oceanic and soil emissions can also contribute to global totals of biogenic VOCs. In general, the following discussion of biogenic VOC emissions will be based on emissions from terrestrial vegetation, as these emissions account for the majority of biogenic VOC emissions.

3.2.4.1

Land vegetation

Trees account for the largest portion of biogenic VOC emissions (Wiedinmyer et al. 2004) and comprise 75% of the total global BVOC (based on the last global estimate of Guenther et al. 1995). The main emission from deciduous trees is isoprene, while terpenoid emissions are dominant from coniferous trees. Certain pine species emit MBO in large quantities equivalent to that of isoprene (Baker et al. 1999; Goldan et al. 1993; Schade and Goldstein 2001). Furthermore, coniferous forests can account for a large fraction of oxVOC emissions,

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including methanol, acetone, ethanol and acetaldehyde (Baker et al. 1999; Goldan et al. 1993; Schade and Goldstein 2001). Both coniferous and deciduous forests can emit VR-BVOC compounds, particularly during the flowering season of deciduous trees (Ciccioli et al. 1999; Holzinger et al. 2004). Other vegetated ecosystems, including shrubs and grasslands, can also be a significant source of biogenic VOC emissions. Despite the limited amount of measurements of isoprene emissions from shrubs, the fraction of shrubs that emit isoprene is generally considered to be similar that of trees (Wiedinmyer et al. 2004). Terpenoid compounds can be emitted in high quantities from shrubs, particularly in Mediterranean ecosystems such as those in California and Europe (Hansen et al. 1997; Winer 1983). Undisturbed grasslands were initially considered to be low emitters (Guenther et al. 1999a; Zimmerman 1979), yet other studies have found that they can emit significant amounts of isoprene, terpenes and oxVOCs (Fukui and Doskey 2000; Koenig et al. 1995). Cutting and drying of grasses can release a wide range of oxVOCs, including oxygenated species and esters (DeGouw et al. 2000; Kirstine et al. 1998; Warneke et al. 2002). Additionally, other phyla in the plant kingdom can also emit isoprene, including ferns (Evans et al. 1985) and mosses (Hanson et al. 1999). Anthropogenically managed land cover, such as crops or urban landscapes, can also emit biogenic VOCs into the atmosphere. Agricultural crops are a source of a wide variety of biogenic VOCs (Arey et al. 1991), yet there are very few studies identifying and quantifying the emissions for a significant range of major crops. Like grasslands, the cutting and harvesting of crops such as hay, alfalfa and corn can also produce significant quantities of oxVOC (DeGouw et al. 2000; Karl et al. 2001b). Urban landscapes, including tree and shrub species, also emit a wide range of quantities and species of biogenic VOC compounds. Additionally, the cutting of grasses in urban areas can release a small but significant amount of oxVOCs into the atmosphere (Karl et al. 2001b).

3.2.4.2

Oceans and soils

Oceans are a small source of biogenic VOC when compared to terrestrial vegetation (Guenther et al. 1995), but can be the dominant source for some specific compounds. Several classes of biogenic VOCs are emitted from oceans, including hydrocarbons, oxVOCs, halocarbons and sulphur-containing compounds such as dimethyl sulphide (DMS). The halogenated and sulphur-containing compounds are technically biogenic VOCs, but we will not focus on them in this review as their impacts on atmospheric chemistry are significantly different from the terrestrially dominant isoprenoid and oxVOC compounds. Small quantities of isoprene, oxVOCs and other hydrocarbons have been measured as emissions from seawater (Bonsang et al. 1992; Donahue and Prinn 1990, 1993; Plass-Dulmer et al. 1995). Isoprene is emitted as by-products of phytoplankton activity in the ocean surface layer and deeper ocean or from seaweeds (Bonsang et al. 1992; Broadgate et al. 2004). Global oceanic isoprene emissions have been estimated to range from 0.2 to 1.2 Tg C/year, and a recent study by Palmer and Shaw (2005) estimated the marine source of isoprene to be 0.1 Tg C/year. Hydrocarbons, including C2 –C5 compounds, are emitted from seawater and are estimated to be ∼2.1 Tg C/year (Bonsang et al. 1992; Plass-Dulmer et al. 1995). While the magnitudes may be relatively small, oceanic sources may be important for remote atmospheric chemistry, particularly in the case of oxVOC. The importance of oxVOCs and their role in atmospheric chemistry has been noted in studies by

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Singh et al. (2000, 2001, 2004) and is discussed in greater detail in Chapter 4. There is conflicting evidence as to whether the ocean is a net source or sink for specific VOCs. Few measurements have been performed to determine the direction of fluxes in the ocean. A recent study by Williams et al. (2004) measured atmospheric and seawater mixing ratios of methanol and acetone and found that the ocean was a sink for methanol and a source for acetone, although the authors caution that these flux directions can vary with location, time of year and the transport of anthropogenic emissions. Inverse modelling studies are often used to determine potential sources and sinks (Heikes et al. 2002; Jacob et al. 2002). Based on several atmospheric measurements and modelling studies, it is believed that the ocean is a source of acetaldehyde (Singh et al. 2003), while acetone is sometimes considered a source (Jacob et al. 2002) and sometimes a sink (Singh et al. 2003). For methanol, Singh et al. (2003) determined that the ocean was in near equilibrium and possibly a net sink. Soils are also considered a minor source of biogenic VOCs or a minor sink, depending on the environmental conditions. However, recent studies have found that oxVOCs can be emitted from bare soil (Schade and Goldstein 2001), with factors such as the amount of litter present, the soil moisture and soil temperature impacting the emission rates. Further research is necessary here to determine the magnitudes of these emissions and their significance with respect to terrestrial emissions.

3.2.5

Environmental controls on biogenic VOC

Biogenic VOC emissions are strongly dependent on environmental conditions, making them particularly sensitive to changes in climate. Light and temperature are two important factors controlling the growth and function of plants, and they are the dominant environmental influences on biogenic VOC emissions as well. In this section, we discuss the impact of immediate (light, temperature, moisture and CO2 -mixing ratios) and long-term (phenology) environmental factors on biogenic VOC emissions.

3.2.5.1

Light dependencies

Since the discovery that plants emit isoprene, the dependence of biogenic VOCs on light has been quantified (Rasmussen and Went 1965; Sanadze 1957). Isoprene emission is strongly dependent on photosynthetically active radiation (PAR; the portion of the spectrum (400–700 nm) that activates photosynthesis). The response of isoprene to PAR is hyperbolic, and emissions increase with increasing light until they reach a saturation point. The instantaneous effects of light on isoprene emissions have been parameterised in order to model biogenic VOC fluxes (Evans et al. 1985; Guenther et al. 1991, 1993; Pandis et al. 1991), and Figure 3.2a shows the common form of this hyperbolic behaviour. Longer-term effects of light on isoprene emissions have also been noted, particularly that sunlit leaves tend to have higher isoprene emission capacities than shaded leaves (Harley et al. 1996; Lerdau and Throop 1999; Litvak et al. 1996). The current parameterisations of light on isoprene emissions are discussed further in Section 3.3.3. Many studies have found that monoterpene emissions are dominantly a function of temperature and are not impacted by PAR (Dement et al. 1975; Guenther et al. 1993). Yet, Steinbrecher (1989) found that certain species of Mediterranean oak emit monoterpenes as a function of both light and temperature in significant quantities that rival isoprene

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(a) 1.2

Light activity factor

1.0 0.8 0.6 0.4 Isoprene and monoterpenes Methylbutenol

0.2 0.0 0

200

400

600

800

1000

1200

1400

PAR (micromoles / m2 h) (b)

Temperature activity factor

2.0

1.5

1.0

0.5 Isoprene Monoterpenes/oxVOC Methylbutenol

0.0 280

290

300

310

Temperature (K) Figure 3.2 (a) Light dependencies of biogenic VOC emissions. (b) Temperature dependencies of biogenic VOC emissions.

production. The mechanisms and functions of these two categories of terpene-emitting plants – temperature-dependent only and temperature- and light-dependent emitters – are quite different. Because the emission of light-dependent monoterpenes is similar to that of isoprene, the isoprene light parameterisation has been used to describe light-dependent emissions of monoterpenes with reasonable results (Ciccioli et al. 1999; Figure 3.2a). Another important light-dependent biogenic VOC is MBO. MBO is an oxVOC emitted from a subset of pine species that grow primarily in the forested regions of the western United States. The presence of MBO was first noted by Goldan et al. (1993), who measured high mixing ratios over a lodgepole pine forest in Colorado, and later by Lamanna and Goldstein (2002), over a ponderosa pine forest in the Sierra Nevada. Both studies found that MBO

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emissions can rival isoprene in these regions with respect to their impact on regional photochemistry. Harley et al. (1998) measured the leaf-level MBO response to light and found a similar response curve as that for isoprene (as shown in Figure 3.2a). While not specifically dependent on light, some oxVOCs can be emitted during light to dark transitions. Holzinger et al. (2000) observed bursts of acetaldehyde within minutes of the transition between light and dark environments, although the exact mechanism for this emission is still uncertain. These findings were also confirmed by Karl et al. (2002) and Graus et al. (2004), who also found that oxVOCs such as hexenal and other C6 VOCs can be emitted from poplar in similar light-to-dark transitions. These emissions represent minor sources to the global budget, but may provide clues about plant metabolism, production and release of the oxVOC species.

3.2.5.2

Temperature dependencies

The temperature dependence of biogenic VOC emissions is noted in early experimental work (Dement et al. 1975) and can influence all categories of biogenic VOC emissions. Tingey et al. (1981) first quantified how isoprene emissions increased with increasing temperature and isolated this response from the light effects described above. Subsequent studies (e.g. Guenther 1991, 1993) found that isoprene emissions increased with temperature until an optimum of 35–45◦ C (Guenther et al. 1993; Harley et al. 1996), after which emissions decreased (shown in Figure 3.2b). Initially, it was thought that this temperature drop-off was due to destruction of the isoprene synthase enzyme (Monson et al. 1992), yet recent investigations indicate that this temperature decrease may be a regulatory mechanism (Singsaas and Sharkey 2000). Longer-term effects of temperature on isoprene emissions have also been measured. For example, Sharkey et al. (1999) found that the temperature of the previous 2 days can affect the basal emission rate (BER, the emission rate at a specific reference temperature and PAR) of a particular tree species, and later work by Petron et al. (2001) found that the temperature of the previous several days can alter the temperature optimum as well as the BER. As described in Section 3.2.2.2, non-light-dependent monoterpenes emitters have a different storage and release mechanism than isoprene and light-dependent monoterpene emitters; therefore, their temperature dependencies are quite different. As shown in Figure 3.2b, temperature-dependent monoterpene emitters have emission rates that increase exponentially with increasing temperature. This is because monoterpenes are stored in pools within the plant, and therefore their emission is linked to their volatility and Henry’s law constant. The temperature dependence of oxVOC and some VR-BVOC emissions have been measured for a few plant species and ecosystems. MBO emissions exhibit a similar temperature dependence as that of isoprene (shown in Figure 3.2b). Harley et al. (1998) observed that leaf-level emissions of MBO increase with temperature until they reach an optimum at approximately 35–40◦ C, after which emissions decrease. As with isoprene, longer-term temperature effects have also been noted for emissions of MBO (Gray et al. 2004; Schade and Goldstein 2001). Schade and Goldstein (2001) noted that most of the oxVOCs emitted from a ponderosa pine plantation were exponentially dependent on temperature and they derived temperature response factors for the measured compounds. At this time, oxVOC emissions are typically represented using a terpene-like temperature dependence in global

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and regional emission models, as shown in Figure 3.2b (e.g. Potter et al. 2003; Schade and Goldstein 2001). VR-BVOCs such as sesquiterpenes are expected to also have strong temperature dependence, and analytical techniques to quantify this relationship are currently under development in several laboratories.

3.2.5.3

Moisture dependencies

Monoterpene and some oxVOC emissions have been found to respond to changing moisture conditions. Emissions of monoterpenes generally increase during and following rain events (Helmig et al. 1998a; Schade et al. 1999) or while foliage is wet (Lamb et al. 1985). Other studies have found positive correlations between ambient relative humidity and monoterpene emissions (Dement et al. 1975; Schade et al. 1999). Measurements of acetone fluxes have been shown to increase with increasing humidity, and ethanol varies as a function of the ambient relative humidity, although the observed relationship may have been driven by changes in stomatal conductance as a function of vapour pressure deficit (Schade and Goldstein 2001).

3.2.5.4

CO2 effects

The influence of atmospheric CO2 mixing ratios on biogenic VOC emissions is an interesting topic in the context of increasing mixing ratios in the atmosphere. However, experimental results have often yielded conflicting results, indicating the need for further research. Laboratory measurements found that short-term exposure to elevated CO2 reduced isoprene emissions (Loreto and Sharkey 1990; Monson 1989) or had no effect (Loreto et al. 1996a). However, longer-term exposure to high CO2 conditions has not yielded consistent results. Sharkey et al. (1991) found that elevated CO2 reduced isoprene emissions from aspen trees, while Staudt et al. (2001) found an increase in isoprene emissions in oak trees under high CO2 . Two recent studies performed in the Biosphere II, a large-scale environmentally controlled experimental chamber, also found a negative feedback between isoprene emission and increasing CO2 . Rosenstiel et al. (2003) found isoprene emissions decreased under elevated CO2 , and Pegoraro et al. (2004) confirmed this isoprene reduction when plants were operating at normal conditions. Constable et al. (1999) found that monoterpene emissions were unaffected by increasing CO2 mixing ratios, yet Loreto et al. (2001a) found significant reductions in monoterpene emissions under the presence of doubled CO2 due to decreased enzyme activity. In forest measurements performed in the vicinity of a natural CO2 spring, Rapparini et al. (2004) found that leaves growing under ambient, elevated CO2 conditions were not inhibited, though leaves subject to a spike in CO2 mixing ratios were reduced.

3.2.5.5

Seasonal variations (phenology)

In addition to the short-term responses of emissions to light, temperature and moisture, the seasonal cycle of vegetation can also influence biogenic VOC emissions. The physical, biological and chemical conditions of foliage change over the span of a growing season, and these changes (known as phenological changes) can impact biogenic VOC emissions. Biogenic VOC emissions respond to seasonal factors such as the outbreak, growth, aging and loss of foliage. An obvious example is the flowering of plants (Arey et al. 1991;

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Ciccioli et al. 1999) and the related release of a wide variety of biogenic VOCs, which humans can detect without sophisticated instrumentation. Isoprene emissions change on phenological timescales, particularly due to the growth, aging and senescence of leaves. In temperature and boreal deciduous forests, isoprene emissions begin about 2–4 weeks after budbreak, require an additional 2–6 weeks to reach their maximum emission capacity and then decline rapidly during leaf senescence (Fuentes et al. 1999; Goldstein et al. 1998; Monson et al. 1994). This delayed timing of emissions has been attributed to a minimum temperature threshold (Monson et al. 1994) as well as isoprene synthase activity (Lehning et al. 2001). Another important factor with respect to phenology is that of leaf age. Many coniferous trees retain leaves longer than the typical growing season as defined by deciduous forests, and recent experimental work has indicated that leaf age can be important for some monoterpene and oxVOC emissions. While emissions of monoterpenes were unaffected by leaf age, Loreto et al. (2001a) did notice a greater storage of monoterpenes in primary vs secondary needles. MBO emissions are observed to decrease with increasing leaf age (Gray et al. 2004; Harley et al. 1998; Schade et al. 2000). Other studies indicate that other oxVOCs such as methanol may have decreasing emissions with increasing leaf age (Nemecek-Marshall et al. 1995). In a tropical forest in Amazonia, Kuhn et al. (2004) measured isoprene and monoterpene fluxes and found that despite little change in biomass and instantaneous environmental conditions, there were large changes in emissions between the wet and dry seasons. They attribute this change in emissions to possibilities such as leaf maturation, changes in leaf growth environment, variable water and nutrient availability, and oxidative capacity of the canopy air. These observations illustrate that we still have much to learn regarding the phenological factors regulating emissions of biogenic VOCs, particularly in tropical environments.

3.3 3.3.1

Emission inventories of biogenic VOCs History of biogenic VOC emission inventories

Since the earliest reports of isoprene emissions from vegetation, there has been significant interest in quantifying the magnitude of biogenic VOC emissions. The first published regional-scale emission inventory was by Rasmussen (1970, 1972), who screened plants to determine the range and magnitude of isoprene emitters. Later measurements and regional inventories were performed by Zimmerman (1979) in the Southeastern United States, Winer (1983) in the state of California, and Lamb et al. (1985) in the northeastern United States. These plant enclosure procedures were used to estimate emission rates for various tree species and implemented with foliar biomass data to calculate biogenic emissions over regional scales. These studies laid the groundwork for estimating biogenic emissions over larger spatial scales. Continental-scale estimates were then published for the United States (Lamb et al. 1987), and a detailed model known as BEIS (Biogenic Emissions Inventory System) was developed for use with regional air quality models (Geron et al. 1994; Pierce and Waldruff 1991). Similar types of inventories have been developed for Great Britain (Stewart et al. 2003), Europe (Simpson et al. 1995) and East Asia (Klinger et al. 2002; Steiner et al. 2002). Early

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global estimates were developed in several studies (Allwine 1992; Dignon and Logan 1990; Mueller 1992), but the most comprehensive global model was developed by Guenther et al. (1995). These early biogenic VOC inventories were focused on isoprene and monoterpene emissions, with the addition of a separate ‘other VOC’ category. The Guenther et al. (1995) model is the most widely cited for global emissions estimates. This model estimates biogenic VOC emissions over large spatial scales, using light and temperature parameterisations of emissions based on leaf- and branch-level studies along with plant distribution estimates. Although some modifications have been made to the initial model, the core structure remains the same and this is described below in Section 3.3.2.1. Recent work has focused on refining these existing algorithms and including chemical species other than isoprene and monoterpenes. For example, Otter et al. (2003) included light-dependent monoterpene emissions in a regional emissions inventory for southern Africa, and other studies have attempted to quantify global oxVOC estimates for acetone (Jacob et al. 2002; Potter et al. 2003), acetaldehyde (Fall 2003) and methanol (Galbally and Kirstine 2002; Heikes et al. 2002). Table 3.2 delineates the various estimates of global biogenic VOC emissions, and Section 3.3.3 discusses how these inventories have changed over time.

3.3.2

Methodology of estimating emission inventories

At present, there are two different general modelling methods used to estimate biogenic VOC emissions. The first is an empirical method developed by Guenther et al. (1995), hereinafter referred to as the Guenther 95 (G95) method. This is a widely accepted and utilised model in many studies of regional and global atmospheric chemistry. The emissions estimates correspond fairly well with measurements of emissions, although there are many regions of the Earth for which the model has not been tested against observations. The second method is a physiologically based or mechanistic model that can respond to potential changes in the vegetation environment. While the physiologically based models are still unreliable on larger scales, they may provide a useful tool in the future to investigate changing emissions under different scenarios of global change.

3.3.2.1

The G95 method

The G95 global model is based on the experimental measurements of Guenther et al. (1991, 1993) and regional emission models used in the United States (Geron et al. 1994; Pierce and Waldruff 1991). Using measurements made in various ecosystems throughout the world, the model estimates a global distribution of isoprene, monoterpene and other VOC emissions. Emission fluxes (F ) are calculated as follows: F = BER × D × γ

(3.2)

where BER is the basal emission rate (g VOC emitted/g biomass measured at standard light and temperature conditions), D is the foliar density of the vegetation (g biomass/m2 ground area) and γ is an activity factor that accounts for light, temperature and leaf age effects. We describe each of these factors below in brief detail. Foliar density (D). Biogenic VOC emissions are extremely dependent on the land cover; therefore, the geographic distribution of vegetation is a crucial portion of the inventory.

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Table 3.2 A history of global biogenic VOC estimates Category

Global source (Tg C/year)

Year

Reference

432 480 827 491 1 150 312–1 062

1965 1968 1992 1992 1995 2004

Rasmussen and Went 1965 Robinson and Robbins 1968 Allwine 1992 Mueller 1992 Guenther et al. 1995 Wiedinmyer et al. 2004

350 452 450 175 285 420 250 503 559 250–750

1979 1988 1990 1990 1991 1992 1992 1995 2001 2004

Zimmerman 1979 Rasmussen and Khalil 1988 Dignon and Logan 1990 Taylor et al. 1990 Turner et al. 1991 Allwine 1992 Mueller 1992 Guenther et al. 1995 Potter et al. 2001 Wiedinmyer et al. 2004

Monoterpenes

480 143 128 147 127

1979 1990 1992 1992 1995

Zimmerman 1979 Taylor et al. 1990 Allwine 1992 Mueller 1992 Guenther et al. 1995

Total other VOCs

279 94 520

1992 1992 1995

Allwine 1992 Mueller 1992 Guenther et al. 1995

28(19–47) 37.5(14–80)b 100 39–117 48

2000 2002 2002 2002 2004

Singh et al. 2000 Galbally and Kirstine 2002 Heikes et al. 2002 Tie et al. 2003 Singh et al. 2004

Total BVOC

Isoprene

Oxygenated VOCs Methanola

Ethanola

3(2–4)

2004

Singh et al. 2004

Acetonea

6(3–11)c 9(6–12)c 20(15–26)d 33–107e 31(15.5–46.5)c

1994 2000 2002 2003 2004

Singh et al. 1994 Singh et al. 2000 Jacob et al. 2002 Potter et al. 2003 Singh et al. 2004

19(11–27)

2004

Singh et al. 2004

Acetaldehydea

a Values converted to Tg C/year. b Emissions from higher plants only. c Primary biogenic emissions only. d Emissions from terrestrial vegetation (based on the best estimate). e Emissions from live vegetation only.

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Volatile Organic Compounds in the Atmosphere

Typically, the land cover information required includes (a) the type of land cover present and (b) the density and amount of biomass present. The level of detail available for these estimates can vary from urban-scale high-resolution inventories to global-scale coarse-resolution estimates. For example, urban-scale air quality studies in Los Angeles have used highresolution land cover data to estimate emission inventories at the scale of 1 km (e.g. Scott and Benjamin 2003), while the G95 model used a 0.5 degree resolution to determine biogenic VOC emissions at the global scale. Once a land cover database with appropriate resolution has been selected, the amount of biomass can be estimated from a variety of scale-dependent sources. Urban- and local-scale models can use a detailed forestry dataset that includes specific information about the types of species and amount of biomass present (e.g. Geron et al. 1994, for the eastern United States, and Solmon et al. 2004, for France). At a global level, more generalised ecosystem data is utilised (Guenther et al. 1995). The approach for estimating foliar density has changed as ground-based and satellite data describing the Earth’s surface becomes more readily available. Early emissions estimates at the regional scale were usually determined from ground-based, forest-specific data. For example, Lamb et al. (1987) and Geron et al. (1994) used general biomass estimates for D based on the vegetation species type and the class of emitter (i.e. high, low or non-emitter of isoprene). The first global model, G95, calculated D from net primary productivity (determined from algorithms based on the temperature and precipitation per model grid cell; see Lieth 1975) and foliar density estimates by Box (1981). Recent regional scales models for California (e.g. Scott and Benjamin 2003) characterised foliar density as a product of two terms: specific leaf weight (g biomass/m2 leaf area) and the leaf area index (m2 leaf area/m2 ground area). The specific leaf weight is assumed to remain relatively constant by species and is determined from a database of specific leaf weights by vegetation species or ecosystem type. Leaf area index can be estimated at high spatial resolutions by satellites and can capture the seasonal changes in biomass. Basal emission rate. The second key piece of information for models is the amount of biogenic VOCs emitted from various types of vegetation. This is scaled from a term known as the emission factor or BER and is usually expressed as the mass of emissions per mass of leaf biomass per time (e.g. μg c g−1 h−1 ). In the G95 model, it is assumed that each plant has a relatively stable BER that can be determined by a series of measurements on individual plant species. BERs are typically measured by branch or leaf enclosures at a standard temperature and light level, and measurement techniques for determining the BER are described in Section 3.6. Over the past several decades, a large database of isoprene and monoterpene BERs has been developed from measurements taken throughout the world, and these are summarised by Wiedinmyer et al. (2004). However, the range of vegetation species for which BERs have been measured represents only a small fraction of the potential emitters. Other biogenic VOC categories are largely uncharacterised elsewhere on the globe. For other oxVOC and VR-BVOC species, very few measurements have been performed, and large-scale estimates using the G95 modelling approach are not possible at this time. Interestingly, recent studies are finding increasing evidence for the variability of BERs. Early studies noted that longer-term environmental changes can affect the BER on a seasonal scale (Goldstein et al. 1998; Monson et al. 1994) as well as shorter-term changes such as

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temperature conditions on the order of several days (Petron et al. 2001; Sharkey et al. 1999). Geron et al. (2000b) found that the BER changed over the course of a growing season in Quercus alba L. (white oak) leaves, as well as changing up to a factor of two over the diurnal cycle. Funk et al. (2003) found that diurnal BER changes varied between species and Staudt et al. (2004) found that there can be significant intra- and interspecific variability in the BER. All of these findings indicate that the BER can vary over spatial and temporal scales, and this increases uncertainty in the G95 modelling method. Because of the large number of measurements required for determining BERs for species or ecosystems, the possibility of taxonomic relationships to categorise emitters vs nonemitters has been explored as a way to circumvent this problem. In the past, taxonomic categorisation of emission rates has been used (e.g. Guenther 1994; Rasmussen and Khalil 1998) and confirmed vs measurements (He et al. 2004; Karlik and Winer 2001), and this method is currently the only reasonable approach for regions with high numbers of unmeasured species. For example, Harley et al. (2004) screened 125 tree species in Amazonia and separated tree species into ‘isoprene-emitting’ or ‘non-emitting’ categories. They then used a taxonomic method to predict whether an unscreened tree species emits isoprene. However, as noted by Kesselmeier and Staudt (1999), there are limitations to this method. One obvious example is that of the genus Quercus (oak), where some species emit isoprene, some species emit monoterpenes, and others are non-emitters (Harley et al. 1999). Therefore, while this method may be helpful for poorly characterised geographic regions, there is a high potential for error. Activity factors (γ ). The activity factors utilise a standardised BER and correct for instantaneous effects such as light and temperature, as well as a range of longer-term effects such as seasonality, long-term T effects and leaf age. The activity factor is composed of several environmental factors: γ = γL γT γA

(3.3)

where γL is the light correction, γT is the temperature correction, and γA is the leaf age correction (Guenther et al. 2000). The instantaneous effects of light and temperature on emissions are described in detail in Section 3.2.5, and their profiles are shown in Figure 3.2. For light-dependent emissions, this curve is parameterised for isoprene, MBO and lightdependent monoterpene emissions as: αCL Q γL =  1 + α2Q 2

(3.4)

where Q is the amount of incoming light as PAR, and α and CL are empirical coefficients (Guenther et al. 1993, 1995), For light- and temperature-dependent emissions such as isoprene and MBO, emissions increase until a temperature optimum and then decrease: γT =

Eopt CT 2 e(CT 1 x) CT 2 − CT 1 (1 − eCT 2 x )

(3.5)

where x = [(1 − Topt ) − (1/T )]/R, Eopt is the maximum normalised emission capacity, T is the leaf temperature, Topt is the temperature at which Eopt occurs,

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and CT 1 and CT 2 are empirical coefficients based on the activation and deactivation energies (Guenther et al. 1993, 1999b). Temperature-only-dependent emissions, such as monoterpenes and other oxVOCs, are parameterised as follows: γT = eβ(T −Ts )

(3.6)

where β is an empirical coefficient, T is the leaf temperature and Ts is the standard temperature of 303 K. The leaf age factor, γA , was added to the algorithm by Guenther et al. (1999b) and defines Eopt and Topt of equation (1988) as a function of the previous days’ temperature. These factors account for the instantaneous impact of environmental conditions, although it is thought that longer-term effects of light, temperature or leaf age could play an important role. Geron et al. (2000b) and Petron et al. (2001) added an additional factor to Equation (3.5) to account for the longer-term effects of temperature on emissions.

3.3.2.2

Process-based modelling

For estimating the impacts of climate change, it is unclear whether an empirical model such as G95 will be able to accurately assess the effects of a changing environment on biogenic VOC emissions. Several recent studies have attempted to develop a physiologically based method for estimating biogenic VOC emissions. To date, these models have focused on isoprene emissions and the biochemical or biophysical pathway of isoprene production. At present, there are three different process-based models based on recent discoveries of the biochemical pathway of isoprene production. While these models are not widely accepted or applied at the global scale, they warrant mention because they may provide a pathway for future biogenic VOC modelling under different scenarios of climate change. Niinemets et al. (1999) developed an isoprene emission model based on the energy requirements for isoprene synthesis, which links the photosynthetic electron transport rate to isoprene emission rates. A second process-based modelling approach is that of Zimmer et al. (2000), which models the biosynthetic process of isoprene with the MEP pathway (see Section 3.2.2.1 and Figure 3.1). A detailed photosynthesis model is used along with kinetic parameters for each synthesis step, based on Michaelis–Menten enzyme kinetics. Martin et al. (2000) developed the third process-based model that is also grounded in the MEP pathway for isoprene synthesis. The model determines isoprene emission rates based on the three potentially limiting steps of synthesis: the pyruvate (or carbon) supply, the ATP or energy supply for phosphorylation to DMAPP, and the rate of isoprene synthesis from DMAPP.

3.3.3

Changing inventories and the current state of knowledge

In early biogenic VOC inventories, rough estimates were made on global or regional scales in attempts to quantify the magnitudes of biogenic VOC emissions. However, as atmospheric chemistry models have become increasingly sophisticated, the speciation and emission distribution of these compounds has become a more important input parameter. Table 3.2 shows the history of global biogenic VOC emissions inventories and their variations over time (Dignon and Logan 1990; Galbally and Kirstine 2002; Guenther et al. 1995; Heikes et al. 2002; Jacob et al. 2002; Mueller 1992; Palmer et al. 2003; Potter et al. 2001, 2003;

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Rasmussen and Khalil 1988; Rasmussen and Went 1965; Robinson and Robbins 1968; Singh et al. 1994, 2000, 2004; Taylor et al. 1990; Tie et al. 2003; Turner et al. 1991; Went 1960; Wiedinmyer et al. 2004; Zimmerman 1979). One of the more notable changes in the past decade is the addition of the biogenic oxVOC species. Work by Singh et al. (2000, 2001) highlights the importance of oxVOC emissions for atmospheric chemistry, and subsequent studies have attempted to constrain atmospheric budgets of acetone, methanol and acetaldehyde. One example is the addition of MBO in the state of California. MBO can be emitted in large quantities and has proven to be important in the modelling of atmospheric chemistry in the regional level (Scott and Benjamin 2003). In response to recent and ongoing research, a change that is likely over the coming decades is the inclusion of VR-BVOC emissions in global inventories. Another new development is the use of satellite-derived products to estimate and validate emissions inventories. Palmer et al. (2003) estimated isoprene emission distributions over North America using satellite-derived formaldehyde columns and compared these emissions with modelled isoprene emissions based on the G95 method. This provides an interesting new method to constrain the global budgets of some biogenic VOC emissions.

3.4 3.4.1

Global distribution of biogenic VOCs Lifetimes

Once emitted to the atmosphere, sinks for biogenic VOCs include dry and/or wet deposition and atmospheric reactions. Biogenic VOCs are quickly oxidised by OH, O3 and/or NO3 (and occasionally chlorine atoms). The presence of double bonds in the chemical structure of biogenic VOCs causes high tropospheric reactivity and makes chemical reactions their dominant loss mechanism. Table 3.3 lists the calculated atmospheric lifetimes for many biogenic VOC species derived from laboratory-measured reaction rate constants and typical mixing ratios of OH, O3 and NO3 (Atkinson and Arey 2003; Atkinson et al. 1990, 1995, 1999; Calvert et al. 2000; Corchnoy and Atkinson 1990; Grosjean and Grosjean 1994; Meyrahn et al. 1986; Papagni et al. 2001; Reissell et al. 2001; Rudich et al. 1996; Smith et al. 1996). Isoprene lifetimes with respect to OH and NO3 are on the scale of hours and with O3 on the scale of days. Monoterpenes generally react more quickly than isoprene, with lifetimes with respect to OH and NO3 on the scale of minutes to hours and with O3 on the scale of minutes to days. Similar to monoterpenes, sesquiterpenes generally have lifetimes of minutes to hours. OxVOCs have much more variable lifetimes, ranging from minutes for compounds such as linalool to days for a compound such as acetone. VR-BVOCs have lifetimes that are estimated to be on the order of minutes or less. Because of their high reactivity, biogenic emissions can be important sinks for radicals in the atmosphere (Fehsenfeld et al. 1992). Isoprene and its oxidation products are generally found to be the dominant sink for OH over forested regions, accounting for up to 15–60% of the daytime removal of OH, depending on location (Carlslaw et al. 2000; Lamanna and Goldstein 2002; Millet et al. 2005). Additionally, recent studies have shown that in areas under urban influence, nighttime NO3 chemistry can play an important role in ozone formation, and biogenics can act to reduce NO3 mixing ratios as well (Warneke et al. 2002).

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Table 3.3 Calculated atmospheric lifetimes of biogenic volatile organic compounds Lifetimea for reaction with

Biogenic VOC

O3c

OHb

NO3d

Isoprene

1.4 h

1.3 days

1.6 h

Monoterpenes Camphene 2-Carene 3-Carene Limonene Myrcene cis-/trans-Ocimene α-Phellandrene β-Phellandrene α-Pinene β-Pinene Sabinene α-Terpinene γ -Terpinene Terpinolene

2.6 h 1.7 h 1.6 h 49 min 39 min 33 min 27 min 50 min 2.6 h 1.8 h 1.2 h 23 min 47 min 37 min

18 days 1.7 h 11 h 2.0 h 50 min 44 min 8 min 8.4 h 4.6 h 1.1 days 4.8 h 1 min 2.8 h 13 min

1.7 h 4 min 7 min 5 min 6 min 3 min 0.9 min 8 min 11 min 27 min 7 min 0.5 min 2 min 0.7 min

Sesquiterpenes β-Caryophyllene α-Cedrene α-Copaene α-Humulene Longifolene

42 min 2.1 h 1.5 h 28 min 2.9 h

2 min 14 h 2.5 h 2 min >33 days

3 min 8 min 4 min 2 min 1.6 h

61 days (Atkinson et al. 1999) 2.5 days (Reissell et al. 2001) 1.0 days (Corchnoy and Atkinson 1990) 1.3 h (Atkinson et al. 1995) 1.8 h (Atkinson et al. 1995) 52 min (Atkinson et al. 1995) 12 days (Atkinson et al. 1999) 2.4 h (Papagni et al. 2001) 53 min (Smith et al. 1996)

>4.5 yearf

>8 year (Atkinson et al. 1999) >300 days (Reissell et al. 2001) 1.5 year (Corchnoy and Atkinson 1990) 4.1 h (Atkinson et al. 1995) 4.5 h (Atkinson et al. 1995) 6 min (Atkinson et al. 1995) 2.0 year (Atkinson et al. 1999) 7.7 days (Rudich et al. 1996) 9 min (Smith et al. 1996)

Oxygenated VOCs Acetonee Camphor 1,8-Cineole cis-3-hexen-1-ol cis-3-hexenyl acetate Linalool Methanol MBO 6-methyl-5-hepten-2-ol

>235 days (Reissell et al. 2001) >110 days (Atkinson et al. 1990) 6.2 h (Atkinson et al. 1995) 7.3 h (Atkinson et al. 1995) 55 min (Atkinson et al. 1995) >4.5 yearf 1.7 days (Grosjean and Grosjean 1994) 1.0 h (Smith et al. 1996)

Source: Reprinted from Atmospheric Environment, 37: S197–219. Copyright (2003), with permission from Elsevier. a From Calvert et al. (2000) unless noted otherwise. b Assumed OH radical concentration: 2.0 × 106 molecules/cm3 , 12-h daytime average. c Assumed O concentration: 7 × 1011 molecules/cm3 , 24-h average. 3 d Assumed NO radical concentration: 2.5 × 108 molecules/cm3 , 12-h nighttime average. 3 e Photolysis will also occur with a calculated photolysis lifetime of ∼60 days for the lower troposphere, July, 40◦ N Meyrahn et al. (1986). f Estimated.

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Similarly, oxVOCs can be important for controlling radical mixing ratios. Lamanna and Goldstein (2002) estimate that MBO accounts for 21% of OH loss over conifers in the Sierra Nevada of California. While biogenic VOCs are normally considered to be sinks for atmospheric radicals, their oxidation in the atmosphere can also provide a new source of radicals. Reactions of ozone with alkenes including terpenes can have high yields of OH radicals (Paulson and Orlando 1996; Paulson et al. 1999). Additionally, photolysis of some oxVOCs such as acetone and acetaldehyde in the upper troposphere can also lead to production of HOx radicals (Wennberg et al. 1998). These impacts are discussed further in Section 3.5.2.

3.4.2

Source distribution

Because of their high reactivity, atmospheric mixing ratios of biogenic VOCs are highly variable and usually concentrated near the emission source. As described in Section 3.3, the source distribution of biogenic VOCs is primarily dependent on the land cover at the Earth’s surface. Emissions are based on the type of vegetation and the amount of biomass present at each location, causing a wide range of temporal and spatial variability in emissions. To give a general idea of the distribution of these source emissions, global isoprene and monoterpene emissions in July (generally considered the global seasonal maximum) are shown in Figures 3.3 and 3.4 as modelled using the G95 algorithm (Guenther et al. 1995). Isoprene is concentrated in regions with large amounts of deciduous forests, such as the southeastern United States, Europe, central Africa, and the tropical regions of Eurasia. Monoterpene emissions, which are considered to be primarily coniferous, are greatest in the western United States, Scandinavia and the forests of Russia. Global distributions of other classes of compounds or VR-BVOCs are still at an early stage of investigation, and thus current global estimates are not available.

3.4.3

Mixing ratio distributions

The mixing ratio distributions in the atmosphere result from the balance between the sources and sinks of biogenic VOC emissions. We have discussed the source distribution of emissions, as well as the reaction with radicals as a major sink for biogenic VOCs. Because of their fast reactivity, transport and deposition processes play minor roles in the global mixing ratio distribution for most biogenic VOCs. Therefore, horizontal surface mixing ratio distributions generally mirror that of the source emissions shown in Figures 3.3 and 3.4. Temporally, different biogenic VOC species have varying diurnal profiles. Isoprene mixing ratios have a strong diurnal cycle with daytime values ranging from 0 to 10 ppb (depending on vegetation sources) peaking at midday, and nighttime values on the order of several to hundreds of ppt. Monoterpenes have a very different diurnal cycle primarily when their emissions are independent of light. Mixing ratios at night can range up to 3 ppb near source regions and are generally higher than daytime because they can accumulate in a shallower nocturnal boundary layer (Biesenthal et al. 1998; Hakola et al. 2000; Lamanna and Goldstein 2002). The diurnal behaviour of oxVOC is highly variable depending on the emission species. Light- and temperature-dependent emissions such as MBO have a similar

0

00

–2

50

0

0

00 –2

00 15

20

0

50

25

–1

–1 12

50

00 10

0–

00

10

0

50 0–

25

50

25

10

0–

0–

10

0

Volatile Organic Compounds in the Atmosphere

0

106

00 50 0– 10 10 00 00 –1 25 0

25 0– 5

25 0

10 0–

0–

10 0

Figure 3.3 Global distribution of isoprene fluxes in July, in mg C/m2 , based on the GEIA database (Guenther et al. 1995).

Figure 3.4 Global distribution of monoterpene fluxes in July, based on the GEIA database (Guenther et al. 1995).

Biogenic VOCs

107

diurnal cycle to that of isoprene, with peak daytime values reaching 2–5 ppb at midday (Schade and Goldstein 2001). Other biogenic oxVOCs such as acetone and acetaldehyde often have daytime values of approximately several ppb, with mixing ratios that are slightly higher during the daytime than in the nighttime. Methanol typically has mixing ratios about a factor of five higher than other oxVOCs in forested regions (Goldan et al. 1995b; Schade and Goldstein 2001). VR-BVOC mixing ratios have been measured although they are often quite low (on the order of several ppt) due to their high reactivity. Recent work by Holzinger et al. (2004) has found that these emissions may in fact be higher than previously assumed, based on the measurements of their oxidation products and their rapid oxidation in the forest canopy. Vertical variations can also affect the spatial mixing ratio distribution, depending on their transport and fate in the atmosphere. Because of their high reactivity, biogenic VOCs generally do not escape the boundary layer. Based on observations of vertical profiles of biogenic VOCs at various sites, mixing ratios of isoprene and monoterpenes generally decrease rapidly with increasing heights (Andronache et al. 1994; Helmig et al. 1998b; Warneke et al. 2001). Many vertical profiles of oxVOC species have now been performed, yet these compound distributions are more complicated than isoprene and terpenes because of the combination of primary and secondary anthropogenic sources in addition to primary and secondary biogenic sources. Vertical distributions over the Atlantic (Singh et al. 2000) showed decreasing mixing ratios with increasing altitude for acetone and methanol, while in the Pacific (Singh et al. 2001), mixing ratios of methanol (0.9 ppb), acetone (350 ppt) and acetaldehyde (60–100 ppt) were relatively constant with height. Scheeren et al. (2003) measured several oxygenated species in and above the boundary layer over the Eastern Mediterranean, and showed that acetone and methanol (approximately several ppb up to 6 km) decrease with increasing height though much less rapidly than isoprene and monoterpenes, as expected due to their longer lifetimes.

3.5

Impact on photooxidants and atmospheric chemistry

The work of Haagen-Smit in the 1950s determined that photochemical smog was produced by a series of photochemical reactions involving hydrocarbons and nitrogen oxides in the presence of sunlight. In the initial attempts at ground-level ozone control in the 1970s, this led to the control of both nitrogen oxide and anthropogenic hydrocarbon emissions. However, research in the 1980s indicated that in the presence of sufficient NOx , biogenic VOCs can play an important role in the formation of ground-level ozone in both rural and urban regions (Chameides et al. 1988; Trainer et al. 1987), because biogenic VOCs tend to be more reactive than their anthropogenic counterparts (Atkinson and Arey 1998) and their emission rates are relatively high compared to anthropogenic VOCs. As research has progressed, the presence of biogenic VOCs has been shown to strongly influence tropospheric ozone and NOx chemistry at the regional and global scale (e.g. Horowitz et al. 1998; Pierce et al. 1998; Poisson et al. 2000). In this section, we discuss the role of biogenic VOCs in tropospheric-gas- and aerosolphase chemistry. This topic has been reviewed extensively (e.g. Atkinson and Arey 1998, 2003; Monson and Holland 2001); thus, we provide only a brief discussion here. We discuss some of the major impacts of biogenic VOC in the troposphere, including their effects on

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Volatile Organic Compounds in the Atmosphere

photooxidants and ground-level ozone, their influence on the NOx distribution, and the ability of their oxidation products to undergo gas-to-particle conversion, leading to the growth of secondary organic aerosols.

3.5.1

Reaction pathways

Depending on the location and time of day, biogenic VOC emissions generally react with OH or NO3 radicals, O3 or the chlorine atom. For most biogenic VOCs, the reaction with oxidants in the atmosphere occurs by either of the following two main mechanisms: 1. the addition of O3 or OH/NO3 radicals to the double carbon bond in the biogenic VOC; 2. the abstraction of an H atom from the hydrogen–carbon bond by OH or NO3 (Finlayson-Pitts and Pitts 2000). Most biogenic VOCs are likely to follow the addition mechanism rather than the abstraction mechanism, with the exception of double carbon bond aldehydes that tend to react via abstraction (Atkinson and Arey 2003). The following discussion describes the cycle of reactions shown in Figure 3.5. Once the biogenic VOC molecule has been oxidised via the addition or abstraction mechanism, Biogenic VOC

OH

R• O2

ROOH

HO2

RO2•

RO2•

NO2 ROONO2 NO

Carbonyl + Alcohol

RONO2

RO•

Decomposition

O2

Isomerisation

Products

Figure 3.5 General biogenic VOC reaction pathway. Reprinted from Atmospheric Environment, 37: S197– 219. Copyright (2003), with permission from Elsevier.

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109

the primary product is an alkyl radical (R• ). The alkyl radical then immediately reacts with O2 to form an alkyl peroxy (RO•2 ) radical. After the alkyl peroxy radical is formed, it can react with NO, NO2 , HO2 or other alkyl peroxys (RO•2 ). Reactions with NO2 and HO2 terminate the reaction sequence by forming peroxy nitrates (ROONO2 ) and peroxides (ROOH), respectively. However, if the reaction sequence continues via reactions with NO and HO2 , they can yield other important reaction products (Seinfeld and Pandis 1998). The NO pathway is important because it can often lead to the formation of NO2 and, hence, ozone (when NO2 is photolysed forming O, then reacting with O2 to form O3 ), while the HO2 pathway is important because it forms relatively stable products and acts as a sink for radicals. Generally, if NO mixing ratios are low (e.g. in a rural environment), the cycle tends towards the HO2 /RO2 reactions (or peroxy-peroxy reactions) that remove radicals from the system and retard ozone production. Reaction of the alkyl peroxy radical with HO2 forms a peroxide (ROOH), which can then be removed via wet/dry deposition, be photolysed (regenerating OH) or react with OH itself. Reactions of RO•2 can form a variety of products including alcohols, aldehydes and ketones. However, if NO is high such as in polluted regions, then peroxy radicals will tend to react with NO or NO2 . Reactions with NO2 form peroxy nitrates (ROONO2 ), which can also act as a reservoir species for NOx as they can be stable at surface-level temperature and pressure conditions. This stability allows them to be transported for long distances to later dissociate. The formation and fate of peroxy nitrates are discussed in greater detail in Chapter 6 of this volume. Reactions of RO•2 and NO can form either an alkyl nitrate (RONO2 ) or an alkoxy radical (RO• ). The alkyl nitrate is a more stable compound and can remove NOx from the cycle. However, if the reaction forms the alkoxy radical (RO• ), it proceeds to isomerise or decompose into a variety of products. This pathway can result in the formation of NO2 , triggering the production of more ozone. The fate of NO reactions (either forming RONO2 or RO) is determined by the ‘branching ratio’, a ratio representing the likelihood of each NO pathway. An example of biogenic VOC oxidation is the reaction of isoprene and the OH radical. There are six different possible hydroxyalkyl radicals that can form from this reaction, and Figure 3.6 shows one possible reaction pathway of isoprene and OH to completion (Atkinson and Arey 2003). In the pathway shown, OH adds to the primary carbon and then reacts with NO to form an alkoxy radical (assuming sufficient NO is present). OH addition on the first or second carbon occurs in approximately 66% of the isoprene–OH reactions and results in the production of methyl vinyl ketone and formaldehyde (Finlayson-Pitts and Pitts 2000). The other third of the reactions involve the OH addition on the third or fourth carbon, producing methacrolein (not shown). The primary products of the isoprene–OH reaction are methyl vinyl ketone, methacrolein and formaldehyde. Another important biogenic VOC oxidation sequence is the α-pinene-ozone reaction, shown in Figure 3.7. This is one of the dominant loss mechanisms for α-pinene and is representative of the type of reactions typically occurring between terpenes and ozone. When ozone adds to the double bond in α-pinene, an ozonide forms that rapidly decomposes to two different Criegee biradicals (Kamens et al. 1999). One of these biradicals (Criegee 1 in Figure 3.7) then reacts to form pinonic acid, norpinonaldehyde and norpinonic acid. The second biradical (Criegee 2) forms pinic acid and several other products. The primary products of this reaction sequence can undergo gas-to-particle conversion and

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Volatile Organic Compounds in the Atmosphere

OH

CH3

CH3

+

C H 2C

•

CH2

CH2

C

H2COH

CH

CH O2 CH3 C

H2COH

CH

• OO CH2

NO

CH3 H2COH

C

CH

ONO2 CH2 CH3 H2COH

C

CH

•

O

+

NO2

CH2

Decomposition

O •

CH2O

+

O2 HCHO

+

CH2 CH Methyl vinyl ketone H3C

C

HO2

Figure 3.6 Isoprene and OH reaction pathway. (With permission from Finlayson-Pitts and Pitts 2000.)

form secondary organic aerosols, as well as form other important gas-phase species such as carbon monoxide, formaldehyde, and OH and HO2 radicals.

3.5.2

Effects on photooxidants

The oxidation of biogenic VOCs can lead to production of tropospheric ozone via the formation of NO2 . NO2 is formed from the reaction of NO and the peroxy radical, and is quickly photolysed after formation: NO2 + hν → NO + O(3 P) M

O(3 P) + O2 −→ O3

(3.7a) (3.7b)

where M represents a third body (usually an N2 or O2 molecule) required for the termolecular reaction of 3.7b. Because the cycle with biogenic VOC regenerates radicals, the presence

Biogenic VOCs

+

111

O3 O O O

O O

O O O Criegee 1

CHO

Criegee 2

HO2 + O

COOH

Pinonic acid

O

CHO

CO OH H2

Norpinonaldehyde

HOOC

+ Other products COOH

Pinic acid

oxidant

O COOH Norpinonic acid Figure 3.7 α-Pinene and ozone reaction pathway. Reprinted from Environmental Science and Technology, 33: 1430–8. Copyright (1999), with permission from American Chemical Society.

of these molecules in conjunction with sufficient NOx can increase the production of ozone in the boundary layer. This general reaction sequence represents the primary source of ozone in the troposphere. Modelling studies in the late 1980s and early 1990s indicated that biogenic VOCs, in particular isoprene, could be extremely important in estimating regional ozone in both rural and urban locations (Chameides et al. 1988; Council 1991; Trainer et al. 1987). Isoprene can strongly influence the ozone production efficiency (OPE), and was found to account for a 50% increase in OPE during the summertime in the northeastern United States (Hirsch et al. 1996). Regional studies in the western United States and Europe have also found that isoprene can contribute significantly to ozone production (Dreyfus et al. 2002; Singh et al. 2004; Tsigaridis and Kanakidou 2002). Global modeling studies have also shown that ozone formation is sensitive to the presence of isoprene. Wang et al. (1998) found that the lack of non-methane hydrocarbons (including

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Volatile Organic Compounds in the Atmosphere

both anthropogenic and biogenic) in a global photochemistry model reduced ozone mixing ratios by 5–20%, with decreased production near the surface and in the upper atmosphere (above 300 mbar) and increased production in the middle troposphere (300–600 mb), and attribute approximately half of this effect to isoprene. Houweling et al. (1998) found that all non-methane hydrocarbon emissions caused a 17% increase in the global mean tropospheric ozone column and that isoprene from biogenic sources were causing about 70% of this increase. One example of the impacts of oxygenated biogenic VOCs is that of Tie et al. (2003). This study included biogenic methanol emissions in a global tropospheric chemistry model and found that adding these emissions slightly increased O3 (1–2%). Besides the impacts on ozone, the presence of biogenic VOC can also influence other photooxidants such as OH and HO2 . Near source regions, the presence of biogenic VOCs and their fast reactions with OH can lead to a significant sink of OH radicals. The methanol study by Tie et al. (2003) also found that biogenic methanol emissions can slightly increase HO2 mixing ratios (1–2%) and decrease OH (1–3%). This decrease in OH due to the addition of biogenic emissions has also been noted in smaller-scale, canopy-level models (Makar et al. 1999). Biogenic VOCs can also have other impacts on HOx cycling in the troposphere. While isoprene can be a sink for OH, the reaction of ozone with biogenic alkenes can also be a source of OH. This provides a potential OH source at times when mixing ratios would otherwise be negligible (e.g. nighttime) and has been found to be account for 10–15% of the total radical production in the southeastern United States (Paulson and Orlando 1996).

3.5.3

Formation of organic nitrates and sequestering of NOx

Global atmospheric chemistry studies in the late 1980s and 1990s suggested that the export of anthropogenic NOx could influence ground-level ozone formation in other regions (Jacob et al. 1993; Liu et al. 1987; Mauzerall et al. 1996). One of the methods of transporting NOx is via the formation of peroxy nitrates. As shown in Figure 3.5, peroxy nitrates are formed by the reaction of an alkyl peroxy radical (RO2 ) with NO2 and have a generalised structure of ROONO2 . These relatively stable products can then be transported long distances to be dissociated in other regions, leading to the formation of ozone far from the original sources of anthropogenic pollution. Horowitz et al. (1998) found that peroxy nitrates formed from isoprene are the principle peroxynitrates around the globe. The topic of peroxy nitrates is discussed in further detail in Chapter 6 of this volume.

3.5.4

Formation of secondary organic aerosols

Aerosols, which are liquid or solid particles suspended in the atmosphere, play an important role in the chemistry and climate of the atmosphere. They can scatter and absorb incoming radiation, affecting the radiative budget at the surface of the Earth, and provide a site for heterogeneous reactions in the atmosphere. Biogenic secondary organic aerosols (SOA) can be formed via two possible methods. The first is when gas-phase oxidation products condense onto existing particles in the atmosphere. This occurs because some oxidation products have low volatilities, allowing them to condense on pre-existing particles while trying to establish equilibrium between the gas and particle phases. The ability of a biogenic

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113

VOC to form SOA depends on its atmospheric abundance, its chemical reactivity and the volatility of its products (Seinfeld and Pandis 1998). The second formation mechanism is via nucleation, or the formation of new particles, although there is no solid evidence at this time that nucleation is a significant source of aerosol to the atmosphere (Kanakidou et al. 2004). Went (1960) first discussed the biogenic VOC contribution to haze observed over forested regions. Subsequent studies in the 1980s and 1990s found that the reaction of monoterpenes with photooxidants could produce large amounts of aerosols in smog chamber experiments (Kamens et al. 1981; Pandis et al. 1991). Later studies attributed monoterpenes and sesquiterpenes as the primary contributors to the formation of secondary organic aerosols (Griffin et al. 1999a; Hoffmann et al. 1997). These initial smog chamber studies concluded that gas-to-particle conversion was not occurring for the oxidation products of isoprene. Biogenically formed aerosols can make up a large fraction of the particulate mass of the atmosphere. Pandis et al. (1991) estimated that monoterpenes could account for a range of 15–50% of the particulate matter in Los Angeles. Global inventories by Andreae and Crutzen (1997) estimated that SOA formation from biogenic VOCs was 30–270 Tg/year, while a subsequent study by Griffin et al. (1999b) used the equilibrium principle in conjunction with smog chamber experiments to estimate an annual total of 18.5 Tg of aerosol from biogenic hydrocarbons. In contrast to previous studies, recent investigations have concluded that isoprene can be a precursor of secondary organic aerosols via a different mechanism than that of monoterpenes. Compounds such as carbonyls and dienes (e.g. isoprene) can undergo conversion via heterogeneous reactions in the presence of an acid catalyst such as sulphuric acid (Jang and Kamens 2001; Limbeck et al. 2003). Claeys et al. (2004a) found that biogenic aerosols in Amazonia contained large amounts of low-volatility polyols, which are likely derived from the reaction of isoprene and OH, accounting for ∼2 Tg of biogenic SOA/year. Another possible formation pathway is the reaction of isoprene with hydrogen peroxide under acidic conditions in the liquid phase (Claeys 2004b). These newly hypothesised chemical mechanisms indicate a growing understanding and the need for further research into biogenic SOA. For further information, we refer the reader to a recent review by Kanakidou et al. (2004) and to Chapter 8 of this volume.

3.5.5

Impacts of oxVOCs on atmospheric chemistry

Biogenic oxVOC emissions play a significant role in tropospheric chemistry because: (a) they can be carriers of reactive nitrogen and sequester nitrogen in the atmosphere, and (b) they can be photolysed, causing significant production of HOx radicals mainly in the upper atmosphere (Jaegle et al. 2001; Singh et al. 2000; Wennberg et al. 1998). Recent measurements in the remote Pacific indicate that the oxVOC species are about twice as abundant as the C2 –C8 hydrocarbons (Singh et al. 2004), indicating their potential magnitude and importance in understanding atmospheric chemistry. Other measurement campaigns over continental regions have found that oxVOC species can comprise a dominant fraction of the total VOC amount and reactivity, and that a large portion of the total oxVOC is likely from biogenic emissions (Goldan et al. 1995a, 1995b; Lamanna and Goldstein 2002; Millet et al. 2005; Riemer et al. 1998).

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Volatile Organic Compounds in the Atmosphere

Besides oxVOCs from primary sources, the oxidation of other biogenic VOCs can create significant amounts of oxygenated products. For example, after biogenic compounds such as isoprene and monoterpenes are oxidised, they can form products such as formaldehyde, acetone and other carbonyl compounds. Globally, the oxidation of isoprene is thought to be the dominant source of formaldehyde. Additionally, secondary formation of acetone by the oxidation of isoprene or MBO can contribute significantly to the global acetone budget.

3.6 3.6.1

Sampling and measurement techniques Mixing ratio measurement methods

Mixing ratios of biogenic VOCs can be measured using a variety of sampling techniques, including sampling air into canisters, collecting air samples on solid adsorbents and using automated in-situ instrumentation. These methods are briefly reviewed here, and a thorough review of recently used methods can be found in Helmig (1999). Samples collected into canisters are generally brought back to a central laboratory facility and analysed following cryocondensation of VOCs and subsequent injection into a gas chromatograph equipped with a capillary column and one of several possible detectors. Typical detectors include the flame ionisation detector (FID), quadrupole mass spectrometer, and photoionisation detector (PID). Due to the poor recovery of some biogenic VOCs from canisters, particularly oxVOCs and many terpenoid compounds, techniques have been developed for collecting air samples directly onto carbon-based solid adsorbent cartridges that are later thermally desorbed for injection into a gas chromatography system. Special considerations are required for measurement of many higher-molecular-weight compounds such as sesquiterpenes, and recent progress has been made in creating methods appropriate to these compounds (Helmig et al. 2003). Automated in-situ instrumentation has been developed and deployed to field sites in order to create more temporally representative observational data sets of VOCs (e.g. Greenberg et al. 1994) and can be coupled to flux measurement systems to measure pairs of simultaneously collected samples from gradient or relaxed eddy accumulation (REA) systems as described below in the flux measurement section (e.g. Goldstein et al. 1995; Lamanna and Goldstein 2002; Schade and Goldstein 2001). Many recent advances have been made in measuring atmospheric VOCs (including biogenics), which are covered in detail in the last chapters of this book. Chapter 8 discusses gas chromatography-isotope ratio mass spectrometry, which is a technique used to measure the isotopic composition of VOC (recently reviewed by Goldstein and Shaw 2003). Chapter 9 of this volume discusses proton transfer reaction mass spectrometry (PTR-MS), which is a fast-response, high-sensitivity measurement with limited selectivity for individual compounds. This method is now widely used for measuring oxVOCs, biogenic VOC fluxes and biogenic VOC oxidation products. Chapter 10 discusses multidimensional gas chromatography, which has been explored but is not yet widely used for biogenic VOC measurements. This technique is able to separate a complex VOC mixture in atmospheric samples, and could be useful for the speciation of biogenic VOCs and their oxidation products.

Biogenic VOCs

3.6.2

115

Flux measurement methods

Several methods are available to measure the emissions of biogenic VOCs from vegetation. Leaf enclosure and branch enclosure are two flux methods described below to determine the emissions from a small, intact sample of plant matter. The next two flux methods, REA and eddy covariance (EC), are micrometeorological methods used to determine fluxes from ecosystems and generally represent fluxes over an area of 105 m2 or more (Cao and Hewitt 1999). The enclosure methods have the advantage of allowing us to understand the contribution of individual species to biogenic emissions and to manipulate the plant environment to determine response to variables such as light and temperature. The micrometeorological methods provide a broader picture of the hydrocarbon fluxes into and out of a whole undisturbed ecosystem, and can be deployed over longer timescales to observe the response of fluxes to changing environmental conditions.

3.6.3

Leaf and branch enclosure measurements

The most direct way to measure biogenic VOC emissions is with leaf and branch chambers. This technique is commonly used to determine emission factors for biogenic VOC emission modelling purposes. Early emissions measurements were taken via branch enclosures (Lamb et al. 1985; Zimmerman 1979), which envelope an entire branch into a Teflon or Tedlar enclosure. Ambient air is then pumped or pulled through the enclosure and an emission rate (ER, mass of emissions per mass of biomass per hour) is determined by the difference in the inlet and outlet air mixing ratios (cin and cout ), the flow rate of the air through the bag, and the biomass within the enclosure: ER =

flow (cout − cin ) mass

(3.8)

ER can also be determined on a per-area basis, by dividing by the leaf area within the enclosure instead of by the mass; however, biogenic VOC emission factors are more typically reported on a per-mass basis (Guenther et al. 1995). Measurements can also be taken at the leaf level, in which a leaf is placed inside a cuvette (chamber). Typically, cuvettes are constructed with the ability to control temperature and light levels to test emissions over a range of conditions and develop emission response curves for modelling purposes. Ambient air is pumped or pulled through the cuvette, and the emission rates are determined in the same manner as Equation (3.8). Past studies have shown that leaf cuvette ER measurements are typically higher than branch enclosure measurements (Guenther et al. 1996), and this has been attributed to light levels and shading within the canopy. When using branch or leaf enclosure techniques, care must be taken not to damage the plant, because that often induces biogenic VOC emissions.

3.6.4

Relaxed eddy accumulation

REA is a micrometeorological method to determine biogenic VOC fluxes in and out of ecosystems or landscapes on a larger scale. This method determines the biosphere–atmosphere flux (F ) of biogenic VOC based on separately collecting and analysing mixing ratios of

116

Volatile Organic Compounds in the Atmosphere

the target species in updrafts (cup ) and downdrafts (cdown ) of air: F = βσw (cup − cdown )

(3.9)

where β is an empirical coefficient that depends on the atmospheric stability, and σw is the standard deviation of the vertical wind speed. A sonic anemometer is used to determine vertical wind speed and direction, and air samples can be collected into canisters or cartridges for later analysis, or can be analysed by automated in-situ instrumentation. For a complete description of this and an example of its application for biogenic VOC flux measurements, see Bowling et al. (1998). This method is particularly effective for measuring ecosystemscale fluxes of compounds that are not amenable to current fast-response measurement methods. For compounds that can be measured with fast time response (better than 1 Hz), the EC method can be used.

3.6.5

Eddy Covariance (EC)

EC is a more direct micrometeorological method to measure biosphere–atmosphere fluxes of biogenic VOCs on the ecosystem scale. Fluxes (F ) are determined from the covariance of the biogenic VOC mixing ratio and the vertical wind speeds using the following: F = Na w  c 

(3.10) w

is deviation from the mean vertical where Na is the number density of air molecules, wind velocity, and c  is the deviation from the mixing ratio of the biogenic VOC. This method requires that the mixing ratio of biogenic VOC be measured as fast or faster than the eddies carry the flux past the sensor or, in other words, faster than the vertical wind is changing directions. Because measurements for the EC method need to be instantaneous, this technique could not be employed for biogenic VOCs until recently. The first published flux measurements using the EC method biogenic VOCs utilised a fast chemiluminescence sensor to measure isoprene flux (Guenther and Hill 1998), and more recent studies have used PTR-MS for monoterpene (Lee et al. 2005; Rinne et al. 2002) and oxVOC fluxes (Karl et al. 2001a; Warneke et al. 2002).

3.7

Future directions

Biogenic VOC emissions are a dominant source of reactive organic gases in the atmosphere. In the 1980s and 1990s, the primary research focus was to quantify emissions and understand their role in atmospheric chemistry, with particular attention on contributions to formation of tropospheric ozone and to the oxVOC budget. However, as measurements of these compounds have progressed, categories of extremely reactive compounds are being found to be biogenic in origin (e.g. oxVOCs, VR-BVOCs). These highly reactive compounds form oxidation products in the gas phase and contribute to growth of secondary organic aerosols, which can play an important role in the interactions between chemistry and climate. The interaction between emission of very reactive compounds and their role in secondary aerosol production and growth are now a dominant focus in biogenic VOC research, and one that we believe will remain prominent in the decade to come.

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Volatile Organic Compounds in the Atmosphere Edited by Ralf Koppmann Copyright © 2007 by Blackwell Publishing Ltd

Chapter 4

Oxygenated Volatile Organic Compounds Ralf Koppmann and Jürgen Wildt

4.1

Introduction

Oxygenated volatile organic compounds (OVOCs) belong to the large family of organic compounds present in the global atmosphere. These compounds include carbonyls, alcohols, ketones, esters, ethers, organic peroxides and organic hydroperoxides. At least the lighter-molecular-weight OVOCs are ubiquitous at relatively high concentrations in the troposphere, and it is widely accepted that they play an important role in atmospheric photochemistry. Despite that recognition, their study has been somewhat neglected, mainly because they are difficult to measure. OVOCs have complex primary and secondary sources. On the one hand they are emitted directly into the atmosphere from a variety of anthropogenic and natural sources. On the other hand, they are products, often the first stable ones, in the gas-phase oxidation pathways of organic compounds in the atmosphere. The removal of OVOCs from the atmosphere occurs mainly by the reaction with OH radicals, but also by photolysis and by wet and dry deposition. However, the removal processes are well understood only for a few compounds. As a consequence, the budgets of OVOCs are either poorly understood or not known at all. While the role of formaldehyde (HCHO) as an oxidation product of methane and other VOCs has been studied for more than two decades, the last years have witnessed a growing interest in the role of other OVOCs owing to improved analytic technologies and a deeper understanding of photochemical processes. As an example, acetone was found to be a key atmospheric compound influencing tropospheric chemistry. It has a significant impact on atmospheric chemistry, especially in the upper troposphere, where photolysis leads to the formation of peroxy acetyl nitrate (PAN). In this way, acetone can influence ozone chemistry by sequestering nitrogen oxides (NOx ) in the form of PAN and by providing free radicals (HOx ) especially in the free troposphere (Jaeglé et al. 2001; McKeen et al. 1997; Singh et al. 1994, 1995; Wennberg et al. 1998). OVOCs may also contribute to organic carbon in aerosols (Jang et al. 2002; Kalberer et al. 2004; Li et al. 2001; Tabazadeh et al. 2004). In addition to their photochemical impact, many of these compounds are harmful to human health especially in urban environments. They affect the respiratory tract and irritate the eyes and some of them are believed to be mutagenic and cancerogenic (Kean et al. 2001; Sin et al. 2001).

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The increasing awareness that these compounds are important players in tropospheric chemistry has lead to significant ongoing research although a quantitative determination of oxygenated hydrocarbons is still a challenge.

4.2

Tropospheric mixing ratios and global distribution

The abundance of OVOCs in the atmosphere has been studied on local, regional, and global scales. The investigated compounds cover a variety of aldehydes, ketones, alcohols, and carboxylic acids. Understandably, the largest set of compounds can be found in urban air, in the immediate vicinity of the sources. In these cases the measurements benefit from high mixing ratios and short transport times from the source to the measurement location. In the context of air quality studies, a large number of investigations have been carried out concerning OVOC concentrations on a local scale. These studies dealt with a large set of aldehydes, ketones, and sometimes alcohols in the urban atmosphere. There are a fewer number of investigations at semi-rural or rural sites, and some studies at remote sites or on a global scale. The investigations on global scales are typically limited to formaldehyde, acetone, methanol, and ethanol. This can be attributed to their ubiquitous production or long lifetimes in the atmosphere leading to larger scale distribution and accessible mixing ratios in remote areas and in the free troposphere, respectively. Table 4.1 gives an overview of these studies and lists the concentrations resp. mixing ratios of the three most abundant OVOCs in the urban atmosphere, formaldehyde, acetaldehyde and acetone. In the following we will give an overview of OVOC measurements sorted by compound groups. Table 4.2 summarises the mixing ratios also for highermolecular-weight OVOCs obtained from field measurements in rural areas from selected studies. Aldehydes. There are numerous studies on atmospheric mixing ratios of formaldehyde and acetaldehyde. A large part of these investigations has been done in cities and populated areas around the world. Formaldehyde and acetaldehyde are the main OVOCs found in urban environments, together contributing to more than half of the OVOCs by mass (Kean et al. 2001; Sin et al. 2001). Typical concentrations of formaldehyde and acetaldehyde observed in urban air range between 1 and 45 μg/m3 (corresponding to 0.8 and 36 ppb) and 0.7 and 35 μg/m3 (corresponding to 0.5 and 19 ppb), respectively. In rural environments mixing ratios are in the lower ppb range, sometimes less than 1 ppb. At remote locations and in the free troposphere mixing ratios seldom exceed 1 ppb. A few studies report observations of higher-molecular-weight aldehydes in urban air. Higher-molecular-weight aldehydes observed in the rural atmosphere often consist of those formed during lipoxygenase (LOX) activity (C6 aldehydes, see section ‘Biogenic emissions’) −C10 aldehydes such as hexanal, heptanal, octanal, nonanal, and decanal. and saturated C6 − These compounds have been found in mixing ratios from <10 ppt up to more than 1 ppb (Ciccioli et al. 1993; Helmig and Greenberg 1994; Wedel et al. 1998; Yokouchi et al. 1990). Emission from vegetation seems to be one of the possible sources. These emissions can be induced by ozone exposure of plants (Karl et al. 2005). The induction can already occur at ozone concentrations below 10 ppb (Wildt et al. 2003). Ozone is present nearly everywhere

Summer 1991 June 1992 Summer 1995 February–May 2000 July – August 1990 February – March 1994 1994, 1995, 1996 September/October 1991 September 1997

January 1991 August 1994

June–July 1990 February 1991 September 1992 May–July 1995 January 1992–December 1995 June 1994/July 1995 July–August 1998 July 1998

Autumn 1984 Summer 1983–1986 December 1987–October 1991 July–August 1992 February–June 1994 November 1993 January–March 1995 March–May 1995 May 1995 May 1997–January 1998 March–August 1999 October 1997–September 2000 July–September 2003

217 – 410 ppt

11.66 ppb

1 ppb–50 ppt 0.6–3.3 ppb

1.2 1.5 1.0 1–8 ppb 2.0

2.07–2.81 3.1–14 2.82–4.81 3.3–3.7 3.2 15–35 ppb 13.7 43.5 3.1–22 1.3–2.8 40 3.7–4.9 13.68

HCHO

0.56 ppb 0.74 ppb 1.56 ppb 134, 202, 166 ppt1

160–40 ppt 0.1–3.8 ppb

1.3 1.4 0.7 1–3 ppb 0.9 0.1–0.6 ppb

1.3 ppb

2.05–2.84 2.3–12 1.78–2.95 4.7–5.7 1.8 15–23 ppb 8.2 28.6 3.6–18 1.1–3.2 — 1.9–2.4 8.33

CH3 CHO

1.8 ppb 1.7 ppb 3.83 ppb 200, 479, 871 ppt1 1.14 ppb 0.2–0.66 ppb 85–2000 ppt 0.1–1 ppb

4.2 ppb 0.7–5.2 ppb 6.2 4.5 3.1 — 4.3

— — — — 2.4 30–70 ppb 10.4 — — — — 1.1–1.5 17.76

CH3 COCH3

Riemer et al. (1998)2 Boudries et al. (2002)2 Singh et al. (1994) Singh et al. (1995)2 Arnold et al. (1997) Reiner et al. (1999)2 Fried et al. (2002)

Arlander et al. (1995a,1995b) Benning and Wahner (1998) and Wedel et al. (1998)2 Lee et al. (1995)

Goldan et al. (1995a)2 Goldan et al. (1995b)2 Slemr et al. (1996) Christensen et al. (2000) Solberg et al. (1996) Müller (1997) Shepson et al. (1991) Grossman et al. (2003)2

Kawamura et al. (2000) Satsumabayashi et al. (1989) Anderson et al. (1996) Grosjean et al. (1993) Granby et al. (1997) Müller (1997)2 Possanzini et al. (2000) Báez et al. (1995) Ferrari et al. (1998) Viskari et al. (2000) Khoder et al. (2000) Sin et al. (2001)2 Feng et al. (2005)2

Reference

Given here are the reported average concentrations or concentration ranges of the three OVOCs (formaldehyde, acetaldehyde and acetone) in μg m−3 unless otherwise noted. In some of these studies other OVOCs have been measured. For these results, we refer to the cited literature. 1 The values refer to measurements in the dark, during the transition period, and during 24-h daylight, respectively. 2 During these studies also other OVOCs have been measured.

SE United States Alert, Canada Free troposphere, United States Pacific Western Europe/Eastern Atlantic Germany North Atlantic

Georgia, United States

Urban sites Los Angeles, United States Takasaki, Japan Denver, United States Atlanta, United States Copenhagen, Denmark Leipzig, Germany Rome, Italy Mexico City, Mexico Grenoble, France Kuopio, Finlannd Cairo, Egypt Hongkong Guangzhou, China Rural sites Alabama, United States Boulder, Colorado, United States Schauinsland, Germany Lille Valby, Denmark Birkenes, Sweden Melpitz, Germany Central Ontario, Canada Pabsthum, NE Germany Remote sites Global Pennewitt, NE Germany

Sampling time

Table 4.1 Overview of studies of OVOCs at urban, rural and remote sites

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Table 4.2 Atmospheric concentrations of OVOC from field measurements in rural areas Compound

Atmospheric concentrations [ppb]

Methanol

6–11

Methanol Methanol Methanol

3.1–22 2–12 ≤7.7

Formaldehyde Formaldehyde Ethanol

480 ng/m3 ≤6.5 0.95–1.2

Ethanol Acetaldehyde Acetaldehyde

0.09–1.32 0.85–1.3 0.4

Acetaldehyde Acetone

360 ng/m3 3.3–4.2

Acetone MBO 1-Penten-3-ol (+methyl-butanals) (Z)-3-Hexenal

1340 ng/m3 2–3 2–6 <0.1–20

Hexenol and Hexanal 1,8 Cineole Hexanal–Decanal Hexanal–Decanal Hexanal–Decanal Methylvinylketone

≤0.02 1.1–8.8 0.7–3.9 ≤0.45 0.4–1.5

Methacrolein

0.5–0.7

Pinonaldehyde

<0.1–7

≤0.15 ppb

Location of measurement

Rural Alabama, Loblolly pine plantation Rural Tennessee Subalpine conifer forest Mixed hardwood forest, northern Michigan Coniferous forest, Hyytiälä Spruce forest, Germany Rural Alabama, Loblolly pine plantation Rural Tenessee Rural Alabama Mixed hardwood forest, northern Michigan Coniferous forest, Hyytiälä Rural Alabama Loblolly pine plantation Coniferous forest, Hyytiälä Sonnblick observatory, Austria Innsbruck, measurements before and after lawn mowing Innsbruck, measurements before and after lawn mowing Maize field, Germany Monti Cimini forest, Italy Rome Maize field, Germany Rural Alabama, Loblolly pine plantation Rural Alabama, Loblolly pine plantation Rural Germany, Waldstein, Norway spruce forest

Reference

Goldan et al. (1995a) Riemer et al. (1998) Baker et al. (2001) Karl et al. (2003) Hellen et al. (2004) Müller et al. (2006) Goldan et al. (1995) Riemer et al. (1998) Goldan et al. (1995a) Karl et al. (2003) Hellen et al. (2004) Goldan et al. (1995a) Hellen et al. (2004) Goldan et al. (1993) Fall et al. (2001) Karl et al. (2001a) Karl et al. (2001a) Wedel et al. (1998) Ciccioli et al. (1993) Ciccioli et al. (1993) Wedel et al. (1998) Goldan (1995a) Goldan (1995a) Müller et al. (2006)

in the lower troposphere at such low concentrations. Emissions of these aldehydes may therefore be nearly permanent. Seasonal variations in the mixing ratios of formaldehyde and acetaldehyde over European rural sites were reported by Solberg et al. (1996). Again, typical mixing ratios ranged between 1 and 4 ppb for formaldehyde and 0.2–1.5 ppb for acetaldehyde depending on the site and season. Summer maxima were observed at all sites, which is in contrast to the seasonal

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133

variations of other VOCs. This may reflect on the importance of primary biogenic emissions over secondary formation due to the oxidation of precursor VOC. Latitudinal and seasonal distributions of formaldehyde were measured over the Pacific and Indian Oceans by Arlander et al. (1990). Mixing ratios varied between 0.7 ppb at mid-northern latitudes and 0.2 ppb at mid-southern latitudes. Meridional cross-sections of the global distribution of formaldehyde and acetaldehyde were measured during the TROPOZ II (Tropospheric Ozone) aircraft campaign covering a latitude range between 70◦ N and 60◦ S, and an altitude range between ground level and about 12 km (Arlander et al. 1995a, 1995b). To our knowledge, these are the only available latitudinal cross-sections of these compounds. For formaldehyde, the data showed a vertical decrease with altitude at all latitudes and a broad maximum around the equator. The measured mixing ratios were considerably higher than those expected from the oxidation of methane alone. Using a 2D model simulation, Arlander et al. could show that the formaldehyde distribution could be explained as soon as they included the formaldehyde formation from the oxidation of −C7 hydrocarbons. The CH3 CHO mixing ratios also showed a decrease with altitude, C2 − but a latitudinal decrease from northern to southern latitudes, in contrast to formaldehyde, obviously as a consequence of higher mixing rations of precursor hydrocarbons in northern latitudes. Figures 4.1 and 4.2 show the 2D distributions of formaldehyde and acetaldehyde, as observed by Arlander et al. Vertical profiles of formaldehyde over the North Atlantic were reported by Fried et al. (2002). Background formaldehyde mixing ratios were around 230 ppt in the free troposphere and decreased between 35◦ N and 50◦ N from about 500 ppt down to 270 ppt in the marine boundary layer. However, they also report elevated mixing ratios of up to 600 ppt in the 4–8 km altitude range. Singh et al. (2000) measured the vertical profiles of formaldehyde, acetaldehyde, and propionaldehyde over the Pacific Ocean. Observed ground-level mixing ratios were about 500 ppt, 400 ppt, and 150 ppt for formaldehyde, acetaldehyde and propionaldehyde, respectively. They report decreasing mixing ratios with altitude (about a factor of four between ground level and 11 km altitude) for all three aldehydes. Except for formaldehyde, GEOSCHEM (Goddard Earth Observing System Chemical)-model-simulated vertical profiles showed much lower mixing ratios than measured, indicating that it is not possible to explain the distribution of these compounds with our current knowledge of atmospheric chemistry. New satellite instruments provide the opportunity to measure global column densities of selected OVOCs. As an example, Figure 4.3 shows column densities of formaldehyde derived from measurement of the SCIAMACHY (Scanning Imaging Absorption Spectrometer for Atmospheric Chartography) instrument aboard ESA’s (European Space Agencies) environmental research satellite (ENVISAT) (Wittrock et al. 2006). Observations of higher-molecular-weight aldehydes are very limited. Typically, mixing ratios are in the range of some 100 ppt down to a few ppt. Gas-phase concentrations of these compounds are discussed in detail in Chapter 9. Alcohols. Several studies show that, in rural areas, methanol is the OVOC with the highest concentration. According to current estimations, methanol seems to be the OVOC with the highest global emissions. Since methanol also exhibits a relatively long lifetime, it is the second most abundant volatile organic compound in the atmosphere. Methanol has anthropogenic and natural sources and can be found in polluted as well as in remote areas.

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Volatile Organic Compounds in the Atmosphere

(a)

10 50

64± 22

55± 21 50± 21

Altitude (km)

8

6

4

220±

214± 35

0

137±

105±

53±

28

31

27

21

55± 21

35±

142±

113±

23

29

39(3) 118± 27

91± 25

151± 39

234±

393±

356± 47

51 314±

690±

116(2)

41

60

195± 33

195± 33

35

200

2

202± 34

13(2)

35 240± 33

54(2)

439± 56

242±

313±

1250±

37

43

154

500

40

400±

20

100

247± 37 191±

135±

33

29

972(2)

1000

374± 44(2)

0

67± 73

20

8 121±

140±

24

25

100

29

141± 29

141±

147±

145± 62(2)

30

193± 33

135± 29

76

118± 36(2)

261±

326±

233±

12

39

44

36

2 394± 71(2)

698± 85

995± 125(2)

713± 95(2)

500

S

229± 21

386± 30

1711±

169±

391±

116±

122(2)

549(4)

146

51

33

0

105± 27

356± 47

534±

1439±

20

77± 26(2)

106± 57(2)

21(3)

100

144± 26

581± 72

1000 300(4)

0 40

211± 35

1031± 696± 545(6) 113(2) 760± 187(2)

53±

117± 28

243± 64(4)

293± 42(4)

51± 21

104± 27

71± 24

376±

527± 86

N

50

207± 34(2)

200

60

60

22

241± 37

187± 33

235± 36

200

40

99± 26

141±

124± 24

6

4

500

56±

64± 25(2)

79± 24

34

335± 66

615± 76

10

199±

140± 30

240± 40

580± 72 1292±

50

Latitude (degrees)

65± 21

Altitude (km)

221±

714± 46

S (b)

19

115±

100

156± 30 215± 41

13±

20

Latitude (degrees)

274± 40

151± 30

171± 66

200

500 40

60 N

Figure 4.1 Meridional cross-sections of the mixing ratios of formaldehyde as observed during the TROPOZ II campaign in January 1991. (a) South-bound flights, (b) North-bound flights. For further details, see Arlander et al. (1995a). With permission from Springer.

Oxygenated Volatile Organic Compounds

135

(a) Altitude (km)

57± 23

4

40± 19

50 93± 28

73± 26

2

63± 45(2)

80

0

60

42± 20

105± 30

80

58± 23

40

75± 26

27± 16 44± 25(2)

50± 22(2)

60± 23

50± 21

20

S

53± 22

60± 23

80

91± 28

41± 22± 10(2) 32(2) 46± 21

0 20 Latitude (degrees)

40

60 N

(b) Altitude (km)

70± 25

4

50± 26

80 2

90± 28

100

0

95± 44(2)

60

40± 26

41± 28

95± 41(2)

112± 114± 46(2) 32(2)

40 S

40± 26

40± 19

100

80

175± 38

123± 131± 73(6) 58(2)

20

131± 43(3)

161± 69(2)

120± 31

106± 30

91± 39(4)

150

177± 39

146± 40

44± 20

104± 49(3)

139± 34

0 20 Latitude (degrees)

40

60 N

Figure 4.2 Meridional cross-sections of the mixing ratios of acetaldehyde as observed during the TROPOZ II campaign in January 1991. (a) South-bound flights, (b) North-bound flights. For further details, see Arlander et al. (1995b). With permission from Springer. 60

HCHO (molecule/cm2)

Latitude

30

2.0 × 1016 1.6 × 1016 1.2 × 1016 8.0 × 1015 4.0 × 1015 0.0

0

−30 −60 −180 −150 −120 −90

−60

−30 0 30 Longitude

60

90

120

150

180

Figure 4.3 The global distribution of formaldehyde derived from SCIAMACHY measurements from August 2004–July 2005 (with permission from F. Wittrock et al. 2006). This image appears in full colour in the plate section that follows page 268 as Plate 2.

In rural areas and over forests, methanol mixing ratios of several ppb are frequently observed. Even values up to 22 ppb have been reported (Riemer et al. 1998). Quite low concentrations of 0.1–1 ppb were measured during a field campaign at the Feldberg, Germany, during August and September 2000 (Folkers 2001). Methanol emissions from plants are attributed to growth processes, and it is therefore not surprising that maximum methanol concentrations in rural areas are observable in spring (Karl et al. 2003).

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Volatile Organic Compounds in the Atmosphere

Compared with methanol, less information exists for ethanol. In rural areas, ethanol mixing ratios range from below 100 ppt to 1.3 ppb (Riemer et al. 1998) and from 100 ppt up to 9.6 ppb (Schade and Goldstein 2001). The main source of ethanol in rural areas is of biogenic origin. Ethanol is emitted from higher plants, particularly under anoxic conditions at the plants roots (Kreuzwieser et al. 1999, 2000). Since these conditions at the roots are extremely variable, the source strength for ethanol and thus atmospheric concentrations will also be variable. The source strength is also highly variable for another group of alcohols. These are the C5 - and C6 -alcohols emitted from plants when membranes are damaged. Examples for this are 1-penten-3-ol or (Z)-3-hexenol. These compounds are synthesised within an enzyme sequence where LOX is a central enzyme. Hence, these compounds are often denominated as LOX products. Since hard stress leading to membrane damage and LOX activity is transient, the emissions are also transient. Such emissions are quite easily detectable during cutting or drying of grass (de Gouw et al. 1999). Atmospheric concentrations therefore occur as pulses with ambient concentrations up to more than 6 ppb for individual LOX products (Karl et al. 2001a). This is quite different for another alcohol, 2-methyl-3-buten-2-ol (MBO). MBO emissions are constitutive and depend on light intensity and temperature in a similar way as isoprene emissions. Emissions are nearly zero during darkness and increase with light intensity. Thus, MBO is observed in concentrations from below measurable up to 7 ppb in the air over strong sources such as Ponderosa pine (Schade and Goldstein 2001). Ketones. The most abundant ketone in the atmosphere is acetone. Except methyl ethyl ketone and methyl vinyl ketone concentrations, which can exceed several ppb, those of all other higher-molecular-weight ketones are lower. For detailed information on these compounds, we refer to Chapter 9 of this book. Local mixing ratios of acetone can be rather high. In polluted areas, values of 10–20 ppb are frequently observed; mixing ratios even up to 70 ppb have been measured. In rural areas, mixing ratios drop to values of several hundred ppt to a few ppb (see Table 4.1). There are a number of studies of acetone concentrations in the free troposphere. Singh et al. (1995) found relatively high concentrations of acetone (0.5 ppb) in the range of 5–10 km at mid-northern latitudes. These concentrations declined to 400 and 200 ppt, respectively, at southern latitudes. Arnold et al. (1997) compiled acetone measurements in the upper troposphere and lower stratosphere, obtained during several aircraft studies at mid-northern latitudes. They report acetone mixing ratios between 200 and 2000 ppt in the upper troposphere (8–13 km) in summer for the different measurement campaigns. During a winter campaign in February 1995, acetone mixing ratios were between 300 and 500 ppt in the upper troposphere, with a steep decrease above the tropopause, which was at about 10 km altitude at that time. Reiner et al. (1999) report vertical profiles of acetone mixing ratios over Germany in September and October 1991. They found mixing ratios between 100 and 1 ppb with a tropospheric average of 830 ppt. Above 10 km, acetone mixing ratios decreased with a scale height of 2.2 km, obviously reflecting increasing photolysis rates and reduced vertical transport in the lower stratosphere. Measured global distribution, such as the ones shown for formaldehyde and acetaldehyde, is not available for acetone. Singh et al. (1995) conducted model studies based on a total

Oxygenated Volatile Organic Compounds

137

acetone source of 49 Tg/year to derive a height–latitude cross-section over the central Pacific at 175◦ E longitude. Carboxylic acids. The current available data on the abundance of carboxylic acids show that the lower-molecular-weight acids, formic acid and acetic acid, are ubiquitous compounds in the troposphere. Higher-molecular-weight carboxylic acids have also been observed in urban, continental and even in the remote troposphere. However, data on these compounds are sparse. Carboxylic acids are found in fog water, cloud water, rain water, snow and ice, in aerosol particles, and in the gas phase (Chebbi and Carlier 1996, and references therein). The highest values of 13 ppb for formic acid and 16 ppb for acetic acid observed so far were reported for Los Angeles air in 1986 (Grosjean 1989). In other urban and rural areas, formic and acetic acid mixing ratios are in the range of several hundred ppt to a few ppb (cf. Grosjean 1989; Hartmann et al. 1989; Kawamura et al. 1985; Lawrence and Koutrakis 1994; Puxbaum et al. 1988). Similar values were observed over the Amazonian rain forest, with the mixing ratios of formic and acetic acid being four times higher in the dry season than in the wet season (Andreae et al. 1988; Talbot et al. 1990). Even in the remote troposphere, mixing ratios up to 1 ppb were observed with pronounced north-south and vertical gradients, obviously reflecting the gradients of the precursor molecules (Arlander et al. 1990; Reiner et al. 1999). The data base for higher-molecular-weight carboxylic acids in the ambient atmosphere is even sparser. There are a few measurements reporting the observation of propionic acid (Kawamura et al. 1985) and pyruvic acid (Andreae et al. 1988), and propionic and n-butyric acid (Satsumabayashi et al. 1989). The mixing ratios were in the range of ten to a few hundred ppt. Dicarboxylic acids are less volatile and are mainly observed in the particulate phase. Kawamura and Kaplan (1990) report that oxalic acid can be observed in the gas phase at elevated temperatures. A compilation of all studies dealing with these compounds can be found in Chebbi and Carlier (1996). In summary, except for some specific compounds such as light aldehydes, methanol and acetone, it is not yet possible to draw a clear picture of OVOCs’ tropospheric distribution and seasonal variations.

4.3

Sources of OVOCs

OVOCs are emitted into the atmosphere from a variety of sources. Anthropogenic sources are automobile exhaust, industrial processes, solvent evaporation and, to a certain extent, biomass burning. On the other hand, the terrestrial biosphere seems to play an important role in emitting these compounds into the atmosphere. As we will see, plants are able to synthesise these compounds and emit them into the atmosphere, even in a much larger variety than anthropogenic sources. All these sources are primary sources, because the OVOCs are emitted directly into the atmosphere. In addition to the emission from primary sources, OVOCs are formed in the atmosphere in the oxidation pathways of other volatile organic compounds. Even during the oxidation of higher-molecular-weight OVOCs, lower-molecular-weight OVOCs are formed. Thus, we have to deal with what we call secondary sources of these compounds. Under ambient conditions, it is often impossible to distinguish between the different origins of these

138

Volatile Organic Compounds in the Atmosphere

compounds. In the following part, we discuss the different primary and secondary sources of OVOCs as well as their contribution to the global budget, and try to summarise the present state of knowledge.

4.3.1

Anthropogenic emissions

Among the industrial sources that have been identified to emit OVOCs directly into the atmosphere are refining and petrochemistry, coal chemistry, paint and varnish industry, sewage treatment and even coffee-roasting (cf. the early but still relevant review by Carlier et al. 1986 and references therein). Saturated alcohols have been used in large quantities as industrial solvents. In the context of air quality, anthropogenic OVOC emissions may have a considerable impact on a local or regional scale. For instance, in Brazil, a nationwide alcohol fuel program has been implemented. Methanol, neat ethanol or ethanol–gasoline mixtures are widely used as fuel for vehicles (de Paula Pereira 1999). Furthermore, OVOCs, especially ethers, are frequently used as fuel additives, such as methyl-tert-butyl ether (MTBE), ethyl-tert-butyl ether (ETBE), tert-butyl alcohol (TBA) and tert-amyl-methyl ether (TAME). A variety of anthropogenic sources is discussed for carboxylic acids (cf. Graedel et al. 1986). Among those are animal waste, combustion of synthetic materials, industrial emissions, waste incineration, automobile exhaust and tobacco smoke. Most of these sources are of local importance. There are only a few estimates of the global source strengths from anthropogenic emissions, for example, for methanol and acetone. A global industrial production of methanol of 24.3 Tg/year was reported by Crocco (1997) for 1996. Based on this figure, a global emission of about 3.7 Tg/year was estimated. This is in agreement with global emission rates of about 4 Tg/year estimated by Singh et al. (1995) based on atmospheric observations. For acetone, Singh et al. (2000) estimate a global emission rate of about 2 Tg/year from anthropogenic sources. These numbers are relatively low compared to other sources. Müller (1992) estimated a global annual emission rate of anthropogenic VOC of about 150 Tg/year. This figure is the sum of about 100 Tg/year from ‘technological sources’ and about 50 Tg/year from ‘biomass burning’ emissions. Although Müller (1992) does not specify the compounds included in this estimate, he states that this figure covers all hydrocarbons from all categories of sources. Since OVOCs are included in this estimation, their global annual emission rate can be expected to be somewhere below this figure. An important contribution to the budget of OVOCs in the atmosphere is due to biomass burning emissions. Biomass burning is associated with deforestation, savannah burning, forest wild fires, agricultural practices and biofuel combustion. Since most of these processes are initiated and controlled by humans, this source is considered as anthropogenic emissions. In the last years, a few studies investigated the emission of OVOCs from this source. Koppmann et al. (1997) reported that the abundance of OVOC in biomass-burning plumes competed with or even exceeded that of other hydrocarbons. Based on this data, they were able to derive emission ratios relative to carbon dioxide for a variety of alcohols, aldehydes, ketones, carboxylic acids and esters. New analytical techniques allowed improving the in-situ measurements of these compounds in biomass burning plumes. For example, Holzinger et al. (1999) used proton

Oxygenated Volatile Organic Compounds

139

transfer reaction mass spectrometry (PTR-MS) to measure oxygenated compounds over Surinam, South America. They report emission ratios relative to carbon monoxide for formaldehyde, acetaldehyde, methanol and acetone. Yokelson et al. (1999) and Goode et al. (2000) identified formic and acetic acid as well as a number of other oxygenated compounds such as methanol and hydroxyethanal in biomass burning emissions using a Fourier transform infrared spectrometer (FTIR) in laboratory and field experiments. Other studies also report significant emissions, mainly of formic and acetic acid (Helas et al. 1992; Lacaux et al. 1993). However, to our knowledge, there are no estimates of source strengths for these compounds available up to now. Emission factors and emission ratios including OVOCs from biomass burning were reviewed by Andreae and Merlet (2001). To our knowledge, there is no specific estimate of the contribution of biomass burning to the global budget of OVOCs available. There are a few estimates of global total VOC emission rates from biomass burning that, in combination with observations, may be used to derive a rough estimate of OVOC emission from this source. The first estimation was made by Crutzen et al. (1979). They calculate a global emission rate of nonmethane hydrocarbons of 28 Tg C/year, with an uncertainty of a factor of three. Lobert et al. (1991) calculated a global emission rate of 42 Tg C/year based on laboratory studies of burning of different types of biomass. Based on field studies of different types of fires, Koppmann et al. (1997) estimated a global VOC emission rate from biomass burning ranging between 25 and 48 Tg (compounds)/year. These figures are in the same range as the figure given by Müller (1992), who based his estimate on the data given by Crutzen et al. (1979), but extended his estimate to other biomass burning contributions. Taking the observation that about 30% of the emitted VOCs are oxygenated compounds (Koppmann et al. 1997), this would add 7.5–15 Tg/year of OVOCs from biomass burning to the total source strength.

4.3.2

Emissions from terrestrial vegetation

Plants are by far the largest natural source of OVOCs in the atmosphere, which warrants a more detailed discussion of the formation and emission mechanisms in a bit more detail here. The compounds emitted are aldehydes, ketones, alcohols and esters. A number of earlier studies have identified OVOCs in the emissions from forests (Isidorov et al. 1985), from agricultural plants (Arey et al. 1991; Winer et al. 1992) and from a variety of agricultural plants, trees and grassland (König et al. 1995). These studies were reviewed by Puxbaum (1997). Saturated and unsaturated alcohols are directly emitted into the atmosphere by vegetation. The emission of methanol from plants was first reported by Fehsenfeld et al. (1992). MacDonald and Kimmerer (1993) observed emissions of methanol from a variety of trees in a similar order of magnitude as isoprene emissions. Since 1993, it is known that 2-methyl-3-buten-2-ol (MBO) is emitted from plants. In a pine forest in Colorado, Goldan et al. (1993) and Harley et al. (1998) observed MBO mixing ratios in the range of several ppb. MBO showed a diurnal cycle as did isoprene; the mixing ratio, however, exceeded that of isoprene by a factor of 5–8. Schade et al. (2000) measured canopy scale fluxes of MBO over a ponderosa pine plantation in California of up to 2 mg C/m2 /h in July and August. On this local scale, this value corresponds to about 0.5% of the gross carbon uptake. MBO thus makes up 20% of the total OH reactivity (Lamanna and Goldstein 1999) and about 50% of the acetone yield (Ferronato et al. 1998).

140

Volatile Organic Compounds in the Atmosphere

A number of compounds, which are commonly referred to as leaf alcohols and leaf esters, have been observed to be present in ambient air. Examples of such compounds are (Z)-3-hexen-1-ol and (Z)-3-hexenylacetate. Additionally, higher-molecular-weight leaf aldehydes such as (Z)-3-hexenal or (E)-2-hexenal are co-emitted with the leaf alcohols. Other alcohols have been observed under various conditions, for example, linalool during the blossoming period of orange trees (Arey et al. 1991). −C10 -OVOCs. Ciccioli et al. Another group of OVOC are saturated aldehydes, mainly C6 − (1993) found these compounds in the atmosphere, and from vertical distributions of the OVOC concentrations they concluded that these OVOCs are emitted from vegetation. Owen et al. (1997) observed direct emissions of hexanal, nonanal and decanal in measurements with plants using a plant enclosure technique. Emissions of these saturated aldehydes were also found from plants under insect attack and from plants exposed to ozone (Karl et al. 2005; Wildt et al. 2003). The main contribution to biogenic emissions occurs in tropical regions, which have a high leaf area index, almost no seasons and more or less constant high temperatures (Guenther et al. 1995). Compared to (see Chapter 3), little is known about the biosynthesis and emission mechanisms of OVOCs. However, direct emissions from plants are known for a couple of OVOCs. Table 4.3 gives an overview of observed emissions for the most abundant OVOCs obtained from field measurements. Table 4.4 summarises results from enclosure and laboratory studies.

4.3.2.1

Biosynthesis of OVOCs

Plants synthesise and emit a large number of volatile organic compounds (Gerzhenzon and Croteau 1991; Knudsen et al. 1993; Puxbaum 1997). These biogenic VOCs are the products of different physiological processes, which take place in different parts of the plants. The biosynthesis of the most abundant biogenically emitted VOCs, isoprene and monoterpenes, has been thoroughly investigated and is well understood (Fall 1999; Kesselmeier and Staudt 1999; Kreuzwieser et al. 1999; Lichtenthaler 1999). Despite a growing number of studies, our knowledge concerning the biosynthesis of OVOCs is still incomplete. In the following, we give an overview of biosynthetic pathways of OVOC formation in the order of increasing numbers of carbon atoms. C1 -OVOCs (methanol, formaldehyde and formic acid). C1 compounds are believed to be synthesised during developmental processes of plants (Kreuzwieser et al. 1999). Meanwhile, it is widely accepted that methanol is synthesised in plants, although – like ethanol – it is toxic in higher concentrations (Joseph and Kelsey 2000). The most probable way of methanol formation seems to be pectin demethylation. Details of this process have been investigated in various studies (Levy and Staehelin 1992; Obendorf et al. 1990; O’Neill et al. 1990; Ricard et al. 1986). Other processes, such as mending of proteins (Mudgett and Clarke 1993) or degradation of lignin (Lewis and Yamamot 1990) seem to be less likely. Methanol emissions have been observed during cell expansion of roots, leaves, and fruits (Fall and Benson 1996), growth and maturation of seeds (Obendorf et al. 1990), cell degradation (Fall and Benson 1996; Levy and Staehelin 1992; Nemecek-Marshall et al. 1995), leaf abscission and senescence of plant tissues (Harrimen et al. 1991) and after exposing tobacco plants to very high ozone concentrations (Beauchamp et al. 2005). Methanol emissions thus seem to be related to growth processes, and also to cell repair. These processes occur both in

Oxygenated Volatile Organic Compounds

141

Table 4.3 Flux densities obtained from field measurements Compound

Estimated flux densities

Location of measurement/plant species

Reference

Baker et al. (2001) Karl et al. (2001b) Schade and Goldstein (2001) Warneke et al. (2002)

Karl et al. (2001b)

Methanol Methanol Methanol Methanol

1 mg/m2 /h 1–8.4 mg/m2 /h 2.9 mg (C)/m2 /h 0.5–1.5 mg/m2 /h

Methanol

0.5 (0.9) mg/m2 /h

Methanol Ethanol Ethanol Acetone Acetone Acetone

0–0.2 mg (C)/m2 /h 0.64 mg (C)/m2 /h 0.17 μg (C)/g(dw)/h 0.37 mg (C)/m2 /h 0.1–1.5 mg/m2 /h 0.5 (0.3) mg/m2 /h

Acetone Acetaldehyde Acetaldehyde Acetaldehyde

0.23 μg (C)/g(dw)/h 0.2 mg (C)/m2 /h 0.5–3 mg/m2 /h Mean 0.3 (0.2) mg/m2 /h

Acetaldehyde Acetaldehyde C5 -LOX-Products + 2-methyl-butanal C6 -LOX-Products

0.5–100 nmol/m2 /min 0.19 μg (C)/g(dw)/h 0.1–0.9 mg/m2 /h

Subalpine conifer forest Hayfield, Austria Ponderosa pine Alfalfa field before and after harvesting Mixed hardwood forest, Michigan Field in Germany Ponderosa pine Norway spruce Ponderosa pine Hayfield, Austria Mixed hardwood forest, Michigan Norway spruce Ponderosa pine Hayfield, Austria Mixed hardwood forest, Michigan Spruce forest, Germany Norway spruce Hayfield, Austria

0.1–1.5 mg/m2 /h

Hayfield, Austria

Karl et al. (2003) Schade and Custer (2004) Schade and Goldstein (2001) Grabmer et al. (2006) Schade and Goldstein (2001) Karl et al. (2001b) Karl et al. (2003) Grabmer et al. (2006) Schade and Goldstein (2001) Karl et al. (2001b) Karl et al. (2003) Cojocariu et al. (2004) Grabmer et al. (2006) Karl et al. (2001b)

aboveground and belowground parts of plants. Due to its high water solubility, methanol can diffuse out of cells and enter the transpiration stream. Hence, methanol produced in roots can be transported to the leaves by the xylem flow and subsequently emitted. Formaldehyde and formic acid are most probably produced in an oxidation process from methanol (Fall 1999). Kesselmeier et al. (1998) report a light-dependent emission either directly caused by a light-dependent process or by the light dependence of stomatal conductance. It is assumed that formic acid is produced from methanol by the enzyme methanol oxidase. The oxidation processes may be dependent on photosynthetic active radiation (PAR). C2 -OVOCs (ethanol, acetaldehyde and acetic acid). Ethanol is synthesised in the plant’s roots due to alcoholic fermentation. With the xylem transportation stream, it is transported to the leaves where oxidation by alcohol dehydrogenase forms acetaldehyde in a first step and acetic acid in a second step (e.g. Kreuzwieser et al. 1999). Possibly, acetic acid is also formed from acetyl coenzyme A (Kesselmeier and Staudt 1999). Biosynthesis and emission of these compounds seem to appear mainly during stress situations, particularly during anaerobic

142

Volatile Organic Compounds in the Atmosphere

Table 4.4 Flux densities obtained from enclosure or laboratory measurements Compound

Fluxes

Methanol

1.5–45 μg/g (dw)/h

Methanol Methanol Formaldehyde Formic acid Ethanol Acetaldehyde

6.8–34 μg/g (dw)/h 0.03–7.6 mg (C)/m2 /h 382–589 ng/g (dw)/h 221–434 ng/g (dw)/h 0.001–0.67 mg (C)/m2 /h 0.3–0.8 mg/m2 /h

Acetaldehyde Acetic acid Acetone Acetone

691–1034 ng/g (dw)/h 36.6–289 ng/g (dw)/h 7.4 ± 1.5 μg/g (fr wt)/h 0.1–0.2 mg/m2 /h

Acetone Butanone Butanone

7.5 × 10−11 mol/m2 /s = 0.06 μg/g (dw)/h 0.1–10.3 mg (C)/m2 /h 11–80 μg/g (dw)

LOX-products

0–10−9 mol/m2 /s

LOX-products (Z)-3-Hexenol

≤23 000 μg (C)/m2 /h ≤1.3 μg/h/g (dw)

(Z)-3-Hexenal + (E)-2-Hexenal Nonanal

30–240 μg/g (dw) ≤0.014 μg/g (dw)/h

Species Eleven different species Russian olive, Aspen Clover, Grass Holm oak and pine Holm oak and pine Clover, Grass Bluegrass, orchard grass Holm oak and pine Holm oak and pine Abies concolor Bluegrass, orchard grass Scots pine Clover Bluegrass, orchard grass Tobacco, sunflower, tomato, Clover and grass Eighteen agricultural and two natural plant species Bluegrass and Orchard grass Sunflower, Scots pine

Reference MacDonald and Fall (1993b) Nemececk-Marshall et al. (1995) Kirstine et al. (1998) Kesselmeier et al. (1997) Kesselmeier et al. (1997) Kirstine et al. (1998) Karl et al. (2001b) Kesselmeier et al. (1997) Kesselmeier et al. (1997) MacDonald and Fall (1993a) Karl et al. (2001b) Shao et al. (2001) Kirstine et al. (1998) Karl et al. (2001b) Heiden et al. (2003) Kirstine et al. (1998) Arey et al. (1991)

Karl et al. (2001a) Wildt et al. (2003)

conditions in the plants’ roots. For example, young poplar trees emit high amounts of acetaldehyde upon flooding of the roots (Kreuzwieser et al. 1999). It is known that the production of ethanol and acetaldehyde also increases as soon as the plant encounters drought stress, extreme cold and high concentrations of ozone or SO2 . Stress-induced emissions are already described by Kimmerer and Kozlowski (1982). Strong increases of acetaldehyde emissions are observed during abrupt light–dark transitions. However, the change of PAR in the environment is much more moderate and insufficient to make this effect important under natural conditions (Graus et al. 2004). C3 -OVOCs. The most important C3 -OVOC known to be emitted from plants is acetone (Fall 1999). The biosynthetic pathways of acetone production are still not known. Acetone is speculated to be synthesised by decarboxylation from acetoacetic acid as an intermediate (Fall 1999; Kesselmeier and Staudt 1999). Another C3 -compound that seems to be directly

Oxygenated Volatile Organic Compounds

143

emitted by plants is propanol. Kirstine et al. (1998) report emissions of propanol from pastures, even though these emissions were low and seemed not to be of large significance. C4 -OVOCs. 2-Butanone is the only a C4 -OVOC that was found to be directly emitted from higher plants (Karl et al. 2001b, Kirstine et al. 1998). Again, not much is known about the biosynthetic pathways leading to the formation of 2-butanone and its emission mechanisms. C5 -OVOCs. Zeidler and Lichtenthaler (2001) have shown that MBO is synthesised from dimethyl-allyl-diphosphate (DMAPP) as is isoprene. Goldan et al. (1993) found very high MBO emissions from conifers in Colorado. They observed mixing ratios of 2–3 ppb in ambient air and diurnal variations with maxima around noon and minima during darkness. The emissions were found to be parallel to those of isoprene, which may be a consequence of similar internal plant formation processes. Since DMAPP is also the precursor of isoprene, it is not surprising that the emissions of MBO are determined by PAR and temperature in a similar way as those of isoprene (Harley et al. 1999). Other C5 -OVOCs emitted from vegetation are 1-penten-3-ol, 1-penten-3-one or 3-pentanone. These C5 -OVOCs are synthesised via the octadecanoid pathway (Fisher et al. 2003). Fall et al. (2001) measured high atmospheric concentrations of C5 -OVOCs after a hard freeze event. These compounds showed concentrations in the range of several ppb over a time period of 5 days and diurnal variations with maximum concentrations during darkness. These compounds belong to the family of LOX products, which will be addressed in the following section. C6 -OVOCs. The octadecanoid pathway is also the origin of C6 -OVOCs, often termed as green leaf volatiles, wound-VOC or LOX products. C6 -OVOCs are well known to most of us since these compounds are emitted upon cell wounding, giving the typical odour during the mowing of grass. The biochemical formation of these compounds is well understood (e.g. Croft et al. 1993; Gardner 1995; Hatanaka 1993). They are formed from the common lipid constituents of plant membranes in a sequence of enzyme reactions. Substrates for these OVOCs are free fatty acids consisting of 18 carbon atoms. These fatty acids are oxidised by LOX enzymes. The hydroperoxides formed in this oxidation process are decomposed by lyase enzymes, and the C6 -OVOCs are formed. Depending on the position where the oxygen is added to the substrate, C9 -OVOCs can also be produced and emitted. Emissions of LOX products are typically observed after wounding (Buttery and Ling 1993; Fall et al. 1999), during drying (de Gouw et al. 1999), after pathogen attack (Croft et al. 1993) and after exposure to ozone at elevated levels (Heiden et al. 1999). The main volatiles produced within this pathway during acute stress situations are the C6 -OVOCs. Emissions of C5 - and C9 -OVOCs are lower than those of the C6 -OVOCs. However, as shown at the example of pathogen-infested tobacco plants, emissions of C5 -LOX products may last longer than those of C6 -LOX products (Heiden et al. 2003). For a given plant species, the emission pattern was found to be independent of the elicitor (Heiden et al. 2003). This behaviour might be understood from the basic processes. Once the free fatty acids are produced, the same processes lead to the emission of these OVOCs, independent of the exact elicitor leading to the production of the free fatty acids. However, except for ozone uptake with very high ozone fluxes (Beauchamp et al. 2005), a definite reason for the emission strength of these OVOC has not yet been identified.

144

Volatile Organic Compounds in the Atmosphere

C10 -OVOCs. Oxygenated monoterpenes can be termed as C10 -OVOCs. Examples are verbenone, myrtenal (C10 H14 O), 1,8-cineole (also known as eucalyptol) or linalool (C10 H18 O). Like MBO, these compounds are also synthesised via DMAPP, implying that they are emitted constitutively. This also implies that the emission strengths may be described with phenomenological algorithms, considering temperature and light intensity as variables determining the emissions. The emission rates of oxygenated monoterpenes seem to be lower than those of non-oxygenated ones. −C10 Saturated aldehydes. Another group of OVOCs are saturated aldehydes (mainly C6 − OVOCs). Owen et al. (1997) showed direct emissions of hexanal, nonanal and decanal in measurements with plants using a plant enclosure technique. These compounds have also been observed in the ambient rural atmosphere (Wedel et al. 1998). Emissions of these OVOCs are also found from plants exposed to ozone or from plants under insect attack (Wildt et al. 2003). It is still unclear whether these aldehydes are synthesised by enzymatic processes or by nonenzymatic oxidative processes inside the plants. Aromatic compounds. Aromatic OVOCs are also emitted by plants. Some of these compounds are synthesised within the phenyl propanoid pathway. Methyl salicylate (C8 H8 O3 ) is the predominant OVOC emitted from this pathway (e.g. Heiden et al. 1999; Seskar et al. 1998). Methyl salicylate is known as a signal molecule for plants (Shulaev et al. 1997); however, there is only one observation of such compounds in the atmosphere known to us (Th. Hoffmann, unpublished results).

4.3.2.2

Emission mechanisms and algorithms

When we discuss OVOC emissions from vegetation, we can distinguish between two classes of OVOCs – OVOCs emitted constitutively and those emitted particularly during stress conditions. Emissions are termed as constitutive emissions if they appear permanently with more or less the same emission strengths, that is the emissions are similar at the same external conditions. Constitutive emissions can be described by phenomenological algorithms. Such algorithms describe the dependence of emission rates on temperature and PAR. The absolute emission rates are considered by standard emission rates often termed as basal emissions or emission factors at fixed temperature and PAR. Phenomenological algorithms are successfully used to describe these dependencies for VOC originating from DMAPP, such as isoprene or monoterpenes (Guenther et al. 1993, 1995; Schuh et al. 1997; Tingey et al. 1991). These algorithms should also be applicable to describe the emission rates of OVOCs originating from DMAPP. This is the case for MBO. To our knowledge, no algorithms have been reported for oxygen-containing monoterpenes. Emissions of oxygenated monoterpenes seem to be lower than those of nonoxygenated ones. For emissions from Quercus ilex and Pinus pinea, Kesselmeier et al. (1997) report that oxygenated monoterpenes contribute between 10 and 26% to the total monoterpene emissions. More data regarding the applicability of phenomenological algorithms to describe the emission rates of other OVOC are sparse. Acetone emissions from pine (Pinus sylvestris L.) depend on temperature as well as on PAR (Shao and Wildt 2002; Shao et al. 2001). These dependencies are also described by a phenomenological algorithm. Kesselmeier et al. (1997) found emissions of formaldehyde, formic acid, acetic acid and acetaldehyde from

Oxygenated Volatile Organic Compounds

Methanol

2.0×10−9

50 1.6×10−9

40

1.2×10−9

30 20

8.0×10−10

10

Temperature (°C)

Φ(Methanol) (mol /m2/s1)

145

4.0×10−10 0 12:00

12:00

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Figure 4.4 Diurnal course of methanol emissions from beech (open squares, left scale). The dotted line indicates the period of illumination. During illumination was 480 μmol/m/s. At day 3, the temperature (black line, right scale) was increased from 20◦ C to 30◦ C, leading to an additional increase in methanol emissions.

Quercus ilex. The diurnal variations of these emissions implied that the emissions are driven by temperature as well as by PAR. The emission rates could be described with the algorithm of Guenther et al. (1993, 1995). However, emissions of formaldehyde and acetaldehyde from pine seemed to be independent of PAR. Also, methanol emissions appear to be permanent and may thus be considered as constitutive emissions. As a typical example, Figure 4.4 shows the diurnal variation of methanol emission from beech measured in a plant chamber under defined conditions. The emission rates clearly show light and temperature dependence. Methanol emissions are related to growth processes. Growth processes may be triggered by temperature or PAR, but they follow plant internal diurnal cycles (Walter and Schurr 2005). Dependencies of methanol emissions on temperature or PAR, therefore, are superimposed by diurnal growth cycles and variations of stomatal aperture. The latter impacts the equilibrium of methanol concentrations between the gas phase in the substomatal cavity and the apoplastic water (Niinemets and Reichstein 2003). From the high water solubility of methanol, it is evident that methanol emissions cannot be described by phenomenological algorithms as they are used today. Emissions of many OVOCs are induced by elicitors such as stress. The impacts of these elicitors vary with time. Hence, such induced emissions are not continuous and cannot be described by phenomenological algorithms developed so far. Only in case of ozone as −C10 -aldehydes (Wildt et al. the elicitor, continuous emissions are observed for the C6 − 2003). The emission rates correlate quite well to each other (except of hexanal in some cases because hexanal emissions may also be due to LOX activity). Also, correlations between their concentrations are observed in the atmosphere (Ciccioli et al. 1993; Wedel et al. 1998). These findings imply similar formation processes and similar emission mechanisms for these OVOCs. The continuity of these emissions allows investigating their dependence on external parameters and, thus, to describe the emissions in dependence of those variables affecting

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the emissions. In case of ozone exposure as an elicitor, temperature and ozone flux dens−C10 -aldehyde emissions. Despite their ity had the most pronounced impacts on the C6 − unknown origin, the emissions of these aldehydes can be described using a phenomenological algorithm. In case of insect or pathogen attack, it is still unknown what measured category can be used as a reference. Thus, in case of these elicitors, an applicable algorithm is missing. The emissions of many other OVOCs are transient. This transient character nearly rules out a quantitative description of stress-induced OVOC emissions.

4.3.3

Natural sources other than vegetation

Besides terrestrial vegetation, other sources also emit OVOCs in non-negligible amounts. Schade and Custer (2004) showed