Smith et al. (2015). Global change pressures on soils from land use and management

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Article (refereed) - postprint

Smith, Pete; House, Joanna I.; Bustamante, Mercedes; Sobocká, Jaroslava;
Harper, Richard; Pan, Genxing; West, Paul C.; Clark, Joanna M.; Adhya,
Tapan; Rumpel, Cornelia; Paustian, Keith; Kuikman, Peter; Cotrufo, M.
Francesca; Elliott, Jane A.; McDowell, Richard; Griffiths, Robert I.; Asakawa,
Susumu; Bondeau , Alberte; Jain, Atul K.; Meersmans, Jeroen ; Pugh,
Thomas A.M. 2016.

Global change pressures on soils from land use and

management

Global Change Biology

, 22 (3). 1008-1028.

10.1111/gcb.13068

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Received Date : 14-Jun-2015

Accepted Date : 17-Aug-2015

Article type : Research Review

Global Change Pressures on Soils from Land Use and Management

Pete Smith

, Jo I. House

, Mercedes Bustamante

, Jaroslava Sobocká

, Richard Harper

Genxing Pan

, Paul West

, Jo Clark

, Tapan Adhya

Cornelia Rumpel

, Keith Paustian

Peter Kuikman

, M. Francesca Cotrufo

, Jane A. Elliott

, Richard McDowell

, Robert I.

Griffiths

, Susumu Asakawa

Alberte Bondeau

, Atul K. Jain

, Jeroen Meersmans

and

Thomas A.M. Pugh

Institute of Biological and Environmental Sciences, Scottish Food Security Alliance-Crops

& ClimateXChange, University of Aberdeen, 23 St Machar Drive, Aberdeen, AB24 3UU,
UK

Cabot Institute, School of Geographical Sciences, University of Bristol, University Road,

Bristol, BS8 1SS, U

Departamento de Ecologia, Universidade de Brasília, I.B. C.P. 04457. Campus

Universitário Darcy Ribeiro - UnB. D.F.. CEP: 70919-970 Brasília, Brazil

National Agriculture and Food Centre Lužianky, Soil Science and Conservation Research

Institute Bratislava, Gagarinova 10, 827 13 Bratislava, Slovakia

School of Veterinary and Life Sciences, Murdoch University, South Street, Murdoch WA.

6150 Australia

Institute of Resources, Environment and Ecosystem of Agriculture, Nanjing Agricultural

University, 1 Weigang, Nanjing 210095, China

Global Landscapes Initiative, Institute on the Environment (IonE), University of Minnesota,

325 Learning & Environmental Sciences, 1954 Buford Ave, St. Paul, MN 55108, USA

Department of Geography and Environmental Science, School of Archaeology, Geography

and Environmental Science, The University of Reading, Whiteknights, PO Box 227,
Reading, RG6 6AB, UK

School of Biotechnology, KIIT University, Bhubaneswar - 751024, Odisha, India

CNRS, Campus AgroParisTech, Bâtiment EGER, 78850 Thiverval-Grignon, France

Department of Soil and Crop Sciences & Natural Resource Ecology Laboratory, Colorado

State University, Fort Collins, Colorado 80523-1499, USA

Alterra Wageningen UR, PO Box 47, 6700AA Wageningen, The Netherlands

Environment Canada, National Hydrology Research Centre, Saskatoon, Saskatchewan,

S7N 3H5, Canada

AgResearch, Invermay Agricultural Centre, Private Bag 50034, Mosgiel, New Zealand

Centre for Ecology & Hydrology, Maclean Building, Benson Lane, Crowmarsh Gifford

Wallingford, OX10 8BB, UK

Graduate School of Bioagricultural Sciences, Nagoya University, Chikusa, Nagoya 464-

8601, Japan

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Mediterranean Institute of Biodiversity and Ecology (IMBE), Aix-Marseille University,

Centre National de la Recherche Scientifique (CNRS) - Institut de Recherche pour le
Développement (IRD) - Université d’Avignon Pays du Vaucluse (UAPV), F-13545 Aix-
en-Provence Cedex 04, France

Department of Atmospheric Sciences, University of Illinois @ Urbana-Champaign 105 S.

Gregory Street, Urbana, IL 61801, USA

Department of Geography, College of Life and Environmental Sciences,

University of

Exeter, Armory Building, Renes Drive, Exeter EX4 4RJ, UK

Karlsruhe Institute of Technology, Institute of Meteorology and Climate Research /

Atmospheric Environmental Research (IMK-IFU), Kreuzeckbahnstrasse 19, 82467
Garmisch-Partenkirchen, Germany

*Corresponding author

: Prof Pete Smith, Tel: +44 (0)1224 272702, Fax: +44 (0)1224

272703, E-mail: pete.smith@abdn.ac.uk

Running head:

Global Change Pressures on Soils

Keywords:

soil, land use change, land use intensity, nitrogen deposition, sulphur deposition,

heavy metal deposition

Paper type:

Invited Review

Abstract

Soils are subject to varying degrees of direct or indirect human disturbance, constituting a

major global change driver. Factoring out natural from direct and indirect human influence is

not always straightforward, but some human activities have clear impacts. These include land

use change, land management, and land degradation (erosion, compaction, sealing and

salinization). The intensity of land use also exerts a great impact on soils, and soils are also

subject to indirect impacts arising from human activity, such as acid deposition (sulphur and

nitrogen) and heavy metal pollution. In this critical review, we report the state-of-the-art

understanding of these global change pressures on soils, identify knowledge gaps and

research challenges, and highlight actions and policies to minimise adverse environmental

impacts arising from these global change drivers.

Soils are central to considerations of what constitutes sustainable intensification. Therefore,

ensuring that vulnerable and high environmental value soils are considered when protecting

important habitats and ecosystems, will help to reduce the pressure on land from global

change drivers. To ensure that soils are protected as part of wider environmental efforts, a

global soil resilience programme should be considered, to monitor, recover or sustain soil

fertility and function, and to enhance the ecosystem services provided by soils. Soils cannot,

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and should not, be considered in isolation of the ecosystems that they underpin and vice

versa. The role of soils in supporting ecosystems and natural capital needs greater

recognition. The lasting legacy of the International Year of Soils in 2015 should be to put

soils at the centre of policy supporting environmental protection and sustainable

development.

Introduction

2015 is the International Year of Soil. This represents an ideal time to take stock of scientific

knowledge about the changing global pressures that humans are exerting on soils. 2015 is

also the year when policy makers will adopt a new legally-binding climate agreement under

the United Nations Framework Convention on Climate Change (UNFCCC), with individual

countries and businesses making policies and targets on greenhouse gas emissions and

removals. Soils storage and cycling of carbon and nitrogen are part of emissions and

removals from the land sector. Furthermore, 2015 is the year when countries will shape and

adopt a new development agenda that will build on the Millennium Development Goals

(MDGs). With increasing population, issues such as food security, water security, energy

security (including bioenergy production) and sustainable integrated land and resource

management are central to many development research and policy agendas. Soils underpin

the provision of many ecosystem services related to development.

Soils provide multiple ecosystem services, allowing sustained food and fibre production, and

delivering climate regulation, flood regulation, improved air and water quality, reducing soil

erosion, and provide a reservoir for biodiversity (Smith et al. 2015). All soils are subject to

some degree of human disturbance, either directly through land-use and land management, or

indirectly through responses to human-induced global change such as pollution and climate

change. Distinguishing natural from direct and indirect human influence is not always

straightforward (Smith, 2005), but some human activities and their consequences have clear

impacts, and despite large heterogeneity in soil properties and responses, robust scientific

knowledge exists.

Human impacts on soils largely emerge from the need to meet the food, fibre, and fuel

demands of a growing population including an increase in meat consumption as developing

nations become wealthier, the production of biofuels, and increasing areas of urbanization.

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This has led to conversion of natural land to managed land (extensification) and

intensification of agricultural and other management practices on existing land such as

increasing nutrient and water inputs and increasing harvest frequency to increase yields per

hectare.

Land cover or land use change (e.g. from forest or natural grassland to pasture or cropland),

removes biomass, changes vegetation and disturbs soils, leading to loss of soil carbon and

other nutrients, changes in soil properties, and changes to above- and below-ground

biodiversity. Some land cover conversions e.g. reforestation after abandonment of cropland,

can increase both above- and below-ground carbon and nutrients. Land use or land

management that does not result in a change of cover (e.g. forest harvest and regrowth,

increased grazing intensity and intensification of crop production), can potentially result in

degradation of soil properties, depending on the characteristics of the management practices.

Land use change has been accelerated by population increases and migration as food, shelter,

and materials are sought and acquired. It is estimated that humans have directly modified at

least 70 Mkm

, or >50 percent of Earth’s ice-free land area (Hooke et al. 2012). The new

Global Land Cover Share-database (Latham et al., 2014) represents the major land cover

classes defined by the FAO. Croplands and grasslands (including both natural grasslands and

managed grazing lands) each covered 13.0 %. “Tree-covered areas” (i.e. both natural and

managed forests) covered 28%, shrub-covered areas 9.5 %. Artificial surfaces (including

urbanised areas) occupy 1 %. Land degradation can be found in all land cover types.

Degraded land covers approximately 24% of the global land area (35 Mkm

). 23% of

degrading land is under broadleaved forest, 19% under needle-leaved forests and 20-25% on

rangeland (Bai

et al

., 2008).

In this review we report the state-of-the-art understanding, and knowledge gaps concerning

impacts of changes in anthropogenic land use and land management on soils, including

interactions with other anthropogenic global change pressures. We also review actions and

policies that limit the adverse impacts arising from these global change drivers. We make the

case to put soils at the centre of research strategy and policy actions as a legacy of the

International Year of Soils.

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Land use/land cover change

Land cover change has been dominated by deforestation, but also conversion of grasslands to

cropland and grazing land. Deforestation has had the greatest impact on historical soil carbon

change, causing on average around 25% of soil carbon to be lost (Guo & Gifford, 2002;

Murty

et al

., 2002). Soil carbon losses largely stem from oxidation of the organic matter as

well as soil erosion.

Deforestation affected an estimated 13 million hectares per year between 2000 and 2010; net

forest loss was 5.2 million hectares per year (FAO, 2010). Most of this recent deforestation

has taken place in tropical countries (FAO, 2010; Hansen et al., 2013). Over 50% of tropical

forest loss occurred in Brazil and Indonesia, largely driven by a few commodities: timber,

soy, beef, and oil palm (West

et al

., 2014). There has been a reduced rate of deforestation in

some regions over the last decade, most notably Brazil (INPE, 2014), largely because of land

use conservation policies (Soares-Filho

et al

., 2014; Nolte

et al

., 2013) as well as economics.

Most developed countries with temperate and boreal forest ecosystems – and more recently,

countries in the Near East and Asia – are experiencing stable or increasing forest areas in

contrast to the large scale historic deforestation in these regions, with afforestation reported in

Europe, USA, China, Vietnam and India (FAO 2013).

Changes in soil properties can vary markedly with type of land cover change, climate, and

method, extent of vegetation removal (e.g. land clearing, fires, mechanical harvest) and

management post-harvest. For example, West

et al

. (2010) estimated that clearing land in the

tropics generally emits three times the amount of carbon per ton of annual crop production

compared to clearing land in temperate areas. Emissions are particularly high when organic

peatland/wetland soils are drained to enable agriculture as the initial soil carbon is higher, and

drainage results in large losses of carbon as previously anaerobic soils become aerobic,

allowing the organic matter to oxidise

For example, clearing forest on organic soils for palm

oil production in Kalimantan emits nine-times more carbon than clearing on neighbouring

mineral soils (Carlson & Curran, 2013). Impacts of deforestation can be reduced by avoiding

deforestation on organic soils, and on steep slopes prone to erosion.

There is large heterogeneity in soil measurements of carbon, nitrogen, microbes etc., and still

many areas of the world with poor data coverage. Models can be used to fill gaps in spatial

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coverage and look at past and future time periods, but these too give very variable results.

Nevertheless there are some clear signals that can be obtained from meta-analyses of field

data and global model results of land use/land cover change with respect to soil carbon.

2.1. Observations of impacts of land cover change

Table 1 presents the results of different meta-analysis studies across different climatic zones

that compared the impacts of land use changes on SOC (Guo & Gifford 2002; Don

et al.

2011; Poeplau

et al.

2011; Bárcena

et al.

2014; Murty

et al.

2002; Wei

et al.

2014a). Changes

in SOC after the conversion of forests to croplands ranged from -24 to -52% without marked

differences between climatic regions. The conversion of pastures to other uses (tree

plantations and particularly, croplands) also induced decreases in SOC (-10% and -59%,

respectively). On the other hand, the substitution of croplands by other land uses (forest

regrowth, tree plantation, grassland, pasture) resulted in an increase of SOC (+18 to +53%).

In the case of afforestation, soil C increase with time after afforestation, and C sequestration

depends on prior land use, climate and the tree species planted.

Fewer meta-analysis studies are available for changes in soil N with changes in land uses. A

compilation with predominance of data from tropical sites indicated that average loss of 15%

of soil N after conversion of forests to croplands (Murty

et al.

2002). In Australia, N losses

after conversion of native vegetation to perennial pasture and cropland were more than 20%

and 38%, respectively (Dalal

et al.

2013) while in China N loss (0-10 cm depth) was 21%

and 31% after 4 and 50 years after conversion of forests to cropland (Wei

et al.

2014b).

Similarly to what was described for SOC, afforestation in subtropical zone results in a

significant increase of N stocks 50 years after conversion (Li

et al.

2012).

2.2. Modelled impacts of land cover change

Dynamic Global Vegetation Models (DGVMs) are used to look at the combined effects of

land use change, climate, CO

, and in some cases N deposition, on vegetation and soil

properties over time. A few global models include some aspects of forest, grassland or

cropland management (Bondeau

et al.

2007; Lindeskog

et al.

2013; Drewniak

et al.

2013;

Jain

et al.

2005). Most DGVMs do not currently model peatland soils. In Tables 1 and 2, and

Figures 1 and 2, we show impacts of past land cover and management change on soil carbon

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and nitrogen as calculated by three DGVMs: ISAM (Jain

et al.

2013; El-Masri

et al.

2013;

Barman

et al.

2014 a,b); LPJ-GUESS (Smith

et al.

2001;

Lindeskog

et al.

2013); and LPJmL

(Bondeau

et al.

2007). The ISAM and LPJ-GUESS models were run with the HYDE

historical land use change data set (History Database of the Global Environment; Klein

Goldewijk

et al.

2011). ISAM included wood harvest following (Hurtt

et al.

2011). The

LPJmL group combined 3 land use change data sets with the geographic distribution of

global agricultural lands in the year 2000. All models were run with historical climate and

, and additionally N deposition in the ISAM model only as it includes a nitrogen cycle.

The effects of land cover change were isolated by comparing model runs with and without

land use/management (Le Quéré

et al.

2014). Table 2 and Figure 1 show the loss of soil

carbon due to historical land use change from 1860 to 2010 (note there was land use change

causing soil carbon loss prior to 1860 particularly in Europe and central Asia, but results are

not shown as they were not available for all three models). As with the observed data (Table

1) high carbon losses are associated with the conversion of forests to croplands. Figure 2

shows the mineral soil C and N concentration of different land cover types in different

geographic ranges.

Differences between the models are large for some systems and regions due to different land

use change data, different land cover definitions, and different processes included in the

models. For example, soil carbon losses are higher in the LPJmL model (Table 2, Figure 1) in

part due to greater land cover change in their land cover reconstructions, while their boreal

grassland soil carbon is high due to the inclusion of permafrost slowing soil carbon

decomposition (Figure 2). Treatment of management processes turns out to be an important

differentiator. ISAM shows strong decreases of soil carbon in some regions e.g. the southern

Boreal zone (Figure 1) where the inclusion of wood harvest removes carbon and nutrients

from the soil, while increases in soil carbon in parts of the mid.-latitudes are due to regrowth

of forest following abandonment of agricultural land.

In semi-arid to arid regions, LPJ-GUESS and LPJmL show opposite signs of soil carbon

change after conversion of natural land to pastures (Figure 1), primarily because LPJ-GUESS

simulates a greater fraction of woody vegetation than LPJmL in these regions under potential

natural vegetation. Conversion of woody vegetation to pasture slightly increases soil carbon

(see the meta analysis of Guo & Gifford 2002), partly because of boosted productivity and

higher turnover rates adding more C to the soil, while the change from potential natural

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grassland to managed pasture (for which the literature is sparse) results in a soil carbon

decrease in LPJmL Pasture management strategies can have a large influence on the soil

carbon storage (see Section 4.3), and may also be partly be responsible for differences.

Vegetation models are embedded in Earth System Models (ESMs) used to project future

climates under different human activity including different land management. Some

significant differences between future model climate projections stem from the differences in

modeling soil carbon, in particular, the strength of the relationship between increasing

temperatures and the increasing rate of soil carbon decomposition (Q

) causing climate-

carbon feedbacks

via

emissions (Friedlingstein

et al.

2006). A recent intercomparison of

11 ESMs used in the IPCC 5

Assessment Report (Todd-Brown

et al.

2013), found the

estimate of global soil carbon from ESMs ranged from 510 to 3040 PgC across 11 ESMs

compared to an estimate of 890-1600 PgC (95% confidence interval) from the Harmonized

World Soil Data Base (FAO/IIASA/ISRIC/ISSCAS/JRC, 2012), with all models having

difficulty representing the spatial variability of soil carbon at smaller (1 degree) scales

compared to empirical data. In all models NPP and temperature strongly influenced soil

carbon stocks, much more so than in the observational data, and differences between models

was found to be largely due to the representation of NPP and the parameterization of soil

decomposition sub-models. A similar, systematic analysis of DGVMs including

benchmarking with observational data, and careful testing of assumptions and process

representations in these models, making use of the very large number of observations that

have become available in the years since these algorithms were formulated (e.g. Medlyn

et al.

2015), could significantly improve model performance. This, along with better representation

of critical biological and geochemical mechanisms would improve model capability (Todd-

Brown

et al.

2013).

2.3 Drainage and conversion of peatlands/wetlands for agriculture

The organic soils in peatlands/wetlands store vast quantities of carbon which decomposes

rapidly when they are drained for agriculture or commercial forestry, resulting in emissions

of CO

and N

O to the atmosphere (Hooijer

et al

., 2010). Other services, in particular water

storage and biodiversity, are negatively impacted. Drainage increases vulnerability to further

losses through fire. The majority of soil carbon is concentrated in peatlands in the boreal zone

and tropical peatland forests in Southeast Asia. These areas, along with wetlands along the

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banks of rivers, lakes and estuaries have increasingly been developed for croplands/bioenergy

production over recent decades. The FAO emissions database estimates that globally there

are 250 000 km

of drained organic soils under cropland and grassland, with total GHG

emissions (N

O plus CO

) of 0.9 Pg CO

eq yr

-1

in 2010, with the largest contributions from

Asia (0.44 Pg CO

eq yr

-1

) and Europe (0.18 Pg CO

eq yr

-1

; FAOSTAT, 2013; Tubiello

et al

2015). Joosten (2010) estimated that there are >500 000 km

of drained peatlands in the

world, including under forests, with CO

emissions having increased from 1.06 Pg CO

-1

in 1990 to 1.30 Pg CO

-1

in 2008, despite a decreasing trend in developed countries, from

0.65 to 0.49 Pg CO

-1

, primarily due to natural and artificial rewetting of peatlands. In

Southeast Asia, CO

emissions from drained peatlands in 2006 were 0.61 ± 0.25 Pg CO

-1

(Hooijer

et al

., 2010). Conversion of peatlands in Southeast Asia is increasing, particularly

for oil palm expansion, where cleared peatlands typically emit ~9 times more carbon than

neighbouring mineral soils (Carlson & Curran 2013). In China, between 1950 and 2000, 13

000 km

of wetland soils were shifted to cultivated arable lands, which led to a SOC loss of

5.5 Pg CO

, mostly from peatlands in Northeast China and Tibet (Zhang

et al

., 2008).

Soil drainage also affects mineral soils. Meersmans

et al

. (2009) showed that initially poorly

drained valley soils in Belgium have lost significant amount of topsoil SOC (i.e. between ~70

and 150 t CO

-1

over the 1960 – 2006 period), most probably as a consequence of

intensified soil drainage practices for cultivation purposes.

Agricultural management

To meet projected increases in food demand, crop production will need to increase by 70-

110% by 2050 (World Bank, 2008; Royal Society of London, 2009; Tilman

et al

., 2011).

This can be achieved either through further expansion of agricultural land (extensification),

or through intensification of production on existing land. Intensification is widely promoted

as the more sustainable option because of the negative environmental consequences of land

expansion through deforestation and wetland cultivation (Foley

et al

., 2011). For example,

Burney

et al

. (2010) estimate that intensification of production on croplands between 1961

and 2010 avoided the release of 590 PgCO

eq. Increased productivity per unit land area can

be achieved through a variety of management practices, such as fertilization, irrigation and

increased livestock density, but these can lead to adverse consequences for the soil and wider

environment (Tilman

et al

., 2002). Intensifying land use can potentially reduce soil fertility

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(without additional inputs) and its ability to sustain high production, as well as soil resilience

to extreme weather under climate change, pests and biological invasion, environmental

pollutants and other pressures. Some key management practices and consequences are

highlighted below and summarised in Table 3.

3.1 Nutrient management

Cultivation of soils results in a decline in soil nutrients (nutrient mining). Nutrient inputs,

from both natural and synthetic sources, are needed to sustain soil fertility and supply nutrient

requirements for crop production. Nutrient supply can improve plant growth which increases

organic matter returns to the soil, which in turn can improve soil quality (see section 3.5), so

balanced nutrient supply has a positive impact on soils (Smith et al., 2015). Overuse,

however, has negative environmental consequences. Annual global flows of nitrogen and

phosphorus are now more than double natural levels (Matson

et al

., 1997; Smil, 2000; Tilman

et al

., 2002). In China, for example, N input in agriculture in the 2000’s was twice that in

1980’s (State Bureau of Statistics-China, 2005).

Between 50-60% of nutrient inputs remain in agricultural soils after harvest (West

et al

2014) and can enter local, regional, and coastal waters becoming a major source of pollution

such as eutrophication leading to algal blooms (Carpenter

et al

., 1998). In many places

around the world, over-use of synthetic nitrogen fertilizers is causing soil acidification and

increased decomposition of soil organic matter, leading to loss of soil function in over-

fertilized soils (Ju

et al

., 2009; Tian

et al

., 2012).

Use of fertilisers and manures contributes to climate change through their energy intensive

production and inefficient use (Tubiello et al., 2015). Globally, approximately 3-5% of

nitrogen additions are released as nitrous oxide (N

O) to the atmosphere when both direct

(from soils) and indirect (e.g. downstream from nitrate leaching) emissions are considered

(Galloway

et al

., 2004), and N

O has ~300 times the radiative forcing of carbon dioxide

(IPCC, 2007). Recent research indicates that the relationship between nitrogen application

and N

O emissions is non-linear, resulting in an increasing proportion of added N being

emitted, as application rate increases (Philibert

et al

., 2013; Shcherbak

et al

., 2014). China,

India, and the United States account for ~56% of all N

O emissions from croplands, with

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28% from China alone (West

et al

., 2014). Overuse of nitrogen and phosphorus fertilizer can

contribute to eutrophication of water bodies, adversely affecting water quality and

biodiversity (Galloway

et al.

, 2003, 2004, 2008).

Nutrient use-efficiency can be significantly increased, and nitrate losses to water and N

emissions can be reduced, through changes in rate, timing, placement, and type of

application, as well as balancing fertilization (Venterea

et al

., 2012; Snyder

et al

., 2014). It

has been estimated that current levels of global cereal production could be maintained while

decreasing global nitrogen application by 50% (Mueller

et al

., 2014).

3.2 Carbon management: reduced disturbance and organic matter additions

Agricultural soils have the potential to store additional carbon than at present if best

management practices are used (Paustian

et al

., 1997; Smith, 2008; Smith, 2012). Soil

organic matter content of soils can be increased through use of improved crop varieties or

grassland species mixtures with greater root mass or deeper roots (Kell, 2012), improved crop

rotations in which C inputs are increased over a rotation (Burney

et al.

, 2010), greater residue

retention (Wilhelm

et al.

, 2004), and use of cover crops during fallow periods to provide

year-round C inputs (Burney

et al.

, 2010; Poeplau & Don, 2015). Several studies report that

soil carbon increases in croplands under no-till management (West & Post, 2002; Ogle

et al

2005). However, the carbon benefits of no-till may be limited to the top 30cm of soil

(Blanco-Canqui & Lal 2008; Powlson

et al

., 2014). Baker

et al

. (2007) found that total soil

carbon was similar in non-till and conventional systems, suggesting that carbon accumulation

is occurring at different depths in the soil profile under different management schemes. Given

the larger variability in sub-surface horizons and lack of statistical power in most studies,

more research is needed on soil carbon accumulation at depth under different tillage regimes

(Kravchenko & Robertson, 2010).

Adding plant-derived carbon from external sources such as composts and biochar can

increase soil carbon stocks. Composts and biochars are more slowly decomposed compared

to fresh plant residues, with mean residence times several (composts) to 10-100 (biochars)

longer than un-composted organic materials (Ryals

et al.

, 2015; Lehmann

et al.

, 2015).

Recent developments suggest that biochar, from the pyrolysis of crop residues or other

biomass, can consistently increase crop N use efficiency while greatly (over 25%) reducing

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direct N

O emissions from N fertilizers (Liu

et al.

, 2012; Huang

et al.

, 2012), as well as

enhancing soil fertility (Woolf

et al

., 2010).

3.3 Water management

The amount of irrigated croplands has doubled in the last 50 years and now accounts for 70%

of all water use on the planet (Gleick, 2003). While irrigated crops cover 24% of all cropland

area, they account for 34% of all production (Siebert & Döll, 2010). Irrigation is concentrated

in precipitation-limited areas such as India, China, Pakistan, and the USA, which account for

72% of irrigation water use (West

et al

., 2014). Agricultural water-use competes with uses

for human and natural ecosystems exacerbating water stress in dry regions. Increased

irrigation has occurred in many areas of world agriculture due to the increasing frequency of

drought under the climate change (West

et al.

, 2014). Where irrigation increases productivity

(e.g. in drought prone areas), organic carbon inputs to the soils would be expected to

increase, increasing soil organic matter content (section 3.2).

Irrigation can increase soil salinity in dry regions with high salt content in the subsoil

(Ghassemi

et al

., 1995; Setia

et al

., 2011). Where salinization occurs, additional irrigation is

needed to “flush” the salts beyond the root zone of the crops, which can further exacerbate

stress on water resources, particularly when using underground water sources. Saline soils,

which have a high concentration of soluble salts, occupy approximately 3.1% (397 Mha) of

the world’s land area (FAO, 1995). Climate change (need for more frequent irrigation) and

increases in human population (increasing demand for more production) are likely to increase

the extent of saline soils (Rengasamy, 2008). The energy required by plants or soil organisms

to withdraw water from the soil or retain it in cells increases with decreasing osmotic

potential. As soils dry out, the salt concentration in the soil solution increases (decreasing

osmotic potential), so two soils of different texture may have the same electrical conductivity,

but the osmotic potential is lower in the soil with low water content (Setia

et al.

, 2011a;

Chowdhury

et al.

, 2011; Ben-Gal

et al.

, 2009). The accumulation of salts in the root zone has

adverse effects on plant growth activity, not only due to negative osmotic potential of the soil

solution resulting in decreased availability of water to plants, but also ion imbalance and

specific ion toxicity (Chowdhury

et al.

, 2011). Salinity affects microorganisms mainly by

decreasing osmotic potential, which affects a wide variety of metabolic activities and alters

the composition and activity of the microbial community (Chowdhury

et al.

, 2011) and

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thereby soil organic matter decomposition.

In saline soils, SOC content is influenced by two opposing factors: reduced plant inputs

which may decrease SOC, and reduced rates of decomposition (and associated mineralisation

of organic C to CO

) which could increase SOC content if the C input were unchanged.

Using a modified Rothamsted Carbon model (RothC) with a newly-introduced salinity

decomposition rate modifier and a plant input modifier (Setia

et al.

, 2011b, 2012), Setia

et al.

(2013) estimated that, historically, world soils that are currently saline have lost an average of

3.47 t SOC ha

-1

since they became saline. With the extent of saline soils predicted to increase

under the future climate, Setia et al. (2013) estimated that world soils may lose 6.8 Pg SOC

due to salinity by the year 2100. Soil salinization is difficult to reverse, but salt tolerant plant

species could be used to rehabilitate salt affected soils (Setia et al., 2013).

Water efficiency can be improved through management practices that reduce water

requirement and evaporation from the soil (such as adding mulch as groundcover), more

precise irrigation scheduling and rates, fixing leaks in dryland irrigation systems, improved

application technology (e.g., drip irrigation) and use of intermittent irrigation in rice paddies.

Given that water limitation is projected to become even more limiting in several semi-arid

regions, e.g. Sub-Saharan Africa, where the human population will probably increase most in

the future, and climate change impacts are projected to be severe, improved water harvesting

methods, e.g. storage systems, terracing and other methods for collecting and storing runoff,

are required to make best use of the limited water resource.

3.4 Harvest frequency

Approximately 9% of crop production increases from 1961-2007 was from increasing the

harvest frequency (Alexandratos & Bruinsma, 2012). The global harvested area (i.e. counting

each time an area is harvested) increased four times faster than total cropland area between

2000 and 2011 (Ray & Foley, 2013). The fraction of net primary production (NPP) extracted

by humans is increasing (Haberl

et al.

, 2007). Global warming is increasing the total area

suitable for double or even triple cropping in subtropical and warm temperate regions (Liu

., 2013). The increase results from fewer crop failures, fewer fallow years, and an increase

in multi-cropping.

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Increasing harvest frequency can reduce soil quality by e.g. continuously removing soil

nutrients and increasing soil compaction through greater soil traffic, but if legumes are

included in rotations as harvest frequency increases, soil quality could be improved.

Increasing harvest frequency may require increasing pesticide and herbicide use, and

increased use of fertilisers contributing to pollution (section 3.1). The net effect will depend

on the effectiveness of the management practices followed.

3.5 Soil compaction

Soil compaction causes degradation of soil structure by increasing soil bulk density or

decreasing porosity through externally or internally applied loads, as air is displaced from the

pores between the soil grains (McCarthy, 2007; Alakukku, 2012). It is the most important

subtype of physical soil deterioration, covering 68 Mha globally when first mapped in the

1990s (Oldeman et al., 1991). Compaction of agricultural soils often results from heavy

machinery or from animal trampling, so is more likely to occur in intensive agricultural

systems (machinery use and high stocking densities), and affects physical, chemical and

biological properties of soil. Top soil compaction can be reversed and controlled, but when

compaction creates impermeable layers in the subsoil, this is less easily reversed.

Subsoil compaction can disrupt nutrient water flows, which in turn can lead to reduced crop

yields, poorer crop quality and can give rise to increased GHG emissions, water and nutrient

run-off, erosion, reduced biodiversity and reduced groundwater recharge (Batey, 2009).

Where compaction cannot be avoided, mitigation is necessary. Biological approaches to

mitigation include planting deep rooted plants such as agroforestry; chemical methods

include fertilization (to overcome yield penalty, though not to remedy compaction); and

technical measures include machinery in which planting does not coincide with wheel tracks,

wide tyres / reduced tyre pressures to reduce pressure per unit area, and precision farming to

retain the same wheel tracks each year (Hamza & Anderson, 2005).

3.6 Livestock density

Livestock production is projected to increase significantly in order to meet the growing

demand from a growing population and increase in per-capita meat consumption, with total

demand for meat expected to grow by more than 200 Mt by 2050 (Alexandratos & Bruinsma,

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2012). The greatest increases in per-capita consumption are projected to be in developing and

transition countries (Alexandratos & Bruinsma, 2012). Since the 1970s, most increased

livestock production has resulted from intensification: increasing livestock density and

shifting to a greater fraction of livestock raised in industrial conditions (Bouwmann

et al

2006). For example, 76-79% of pork and poultry production is industrialized (Herrero &

Thornton, 2013). Manure, inputs for growing feed, and soil loss from intensively managed

areas can be major sources of water pollution to local and downstream freshwater

ecosystems. Clearing natural ecosystems for new pastures, particularly in arid and semi-arid

regions, typically occurs on low-productivity lands with a much higher risk of soil erosion

and soil carbon/nutrient depletion (Alexandratos & Bruinsma, 2012), and negatively impacts

water storage and biodiversity. The impacts of livestock production are particularly prevalent

for beef production, which has a least an order of magnitude greater impact on land, water,

GHGs, and reactive nitrogen compared to other livestock (Eshel

et al

., 2014; Ripple

et al

2014). Moreover, industrial livestock production had led to an increased use of veterinary

medicines, antibiotics and hormones, posing potential risks to soil, water, ecosystems and

human health. Improved grazing management (e.g. optimised stocking density) can reduce

soil degradation, and thereby maintain and enhance organic matter content (McSherry &

Ritchie, 2013; see sections 3.2 and 4.3), and can reduce soil compaction, thereby increasing

infiltration and water storage and reduce risk of runoff and flooding downstream (Marshall

al.

, 2009).

Other land management

4.1 Forest management

Logging and fire are the major causes of forest degradation in the tropics (Bryan

et al

., 2013).

Logging removes nutrients and negatively affects soil physical properties and nutrient levels

(soil and litter) in tropical (e.g. Olander

et al

., 2005; Villela

et al

., 2006; Alexander, 2012)

and temperate forests (Perez

et al

., 2009). Forest Fires affect many physical, chemical,

mineralogical, and biological soil properties, depending on fire regime (Certini, 2005).

Increased frequency of fires contributes to degradation, and reduces the resilience of the

biomes to natural disturbances. A meta-analysis of 57 publications (Nave

et al

., 2011)

showed that fire caused a significant decrease in soil C (-26%) and N (-22%). Fires reduced

forest floor storage (pool sizes only) by an average of 59% (C) and 50% (N), but the relative

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concentrations of these two elements did not change. Prescribed fires caused smaller

reductions in C and N storage (-46% and -35%) than wildfires (-67% and -69%). These

differences are likely because of lower fuel loads or less extreme weather conditions in

prescribed fires, both factors that result in lower fire intensity. Burned forest floors recovered

their C and N pools in an average of 128 and 103 years, respectively. Among mineral soil

layers, there were no significant changes in C or N storage, but C and N concentrations

declined significantly (-11% and -12%, respectively). Mineral soil C and N concentrations

were significantly reduced in response to wildfires, but not after prescribed burning.

Forest fires produce charcoal, or black carbon, some of which can be preserved over centuries

and millennia in soils. Dissolved black carbon (DBC) from burning of the Brazilian Atlantic

forest continued to be mobilized from the watershed each year in the rainy season, despite the

fact that widespread forest burning ceased in 1973 (Dittmar

et al

., 2012).

A large field study in the Amazon (225 forest plots) on the effects of anthropogenic forest

disturbance (selective logging, fire, and fragmentation) on soil carbon pools showed that the

first 30 cm of the soil pool did not differ between disturbed primary forests and undisturbed

areas of forest, suggesting a resistance to impacts from selective logging and understory fires

(Berenguer

et al

., 2014). As with deforestation, impacts of human disturbances on the soil

carbon are of particular concern in tropical forests located on organic soils and on steep

easily-eroded slopes.

4.2 Shifting cultivation

Shifting cultivation practices, where land is cleared through fire, have been practiced for

thousands of years, but recent increasing demographic pressure has reduced the duration of

the fallow period, affecting the system sustainability. Moreover, especially in Southeast Asia

where urbanisation is expanding in fertile planes, shifting cultivation is practiced in sloping

uplands, which are prone to soil and carbon loss by erosion (Chaplot

et al

., 2005). A review

by Ribeiro Filho

et al

. (2013) reported negative impact on SOC associated with the

conversion stage, modified by the characteristics of the burning. Chop-and-mulch of enriched

fallows appears to be a promising alternative to slash-and-burn, conserving soil bulk density,

and significantly increasing nutrient concentrations and organic matter content compared to

burnt cropland, and a control forest in a study in the Amazon (Comtea

et al

., 2012)

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4.3 Grassland management and dryland degradation

Grasslands, including rangelands, shrublands, pastureland, and cropland sown with pasture

and fodder crops, cover 26% of the global ice-free land area and 70% of the agricultural area,

and contain about 20% of the world’s soil organic carbon (C) stocks. Grasslands on every

continent have been degraded due to human activities, with about 7.5% of grassland having

been degraded because of overgrazing (Conant, 2012). A meta-analysis (McSherry & Ritchie,

2013) of grazer effects on SOC density (17 studies that include grazed and un-grazed plots)

found higher grazing intensity was associated with increased SOC in grasslands dominated

by C4 grasses (increase of SOC by 6–7%), but with lower SOC in grasslands dominated by

C3 grasses (decrease of SOC by an average 18%). An increase in mean annual precipitation

of 600 mm resulted in a 24%

decrease

in the magnitude of the grazer effect on finer textured

soils, but on sandy soils the same increase in precipitation produced a 22%

increase

in the

grazer effect on SOC (McSherry & Ritchie, 2013).

Land use dynamics and climate change are the major drivers of dryland degradation with

important feedbacks, with changes in plant community composition (e.g. shrub encroachment

and decrease in vegetation cover; D’Odorico

et al

., 2013). A review by Ravi

et al

. (2010)

indicated soil erosion as the most widespread form of land degradation in drylands, with wind

and water erosion contributing to 87% of the degraded land. Grazing pressure, loss of

vegetation cover, and the lack of adequate soil conservation practices increase the

susceptibility of these soils to erosion. The degree of plant cover is negatively related to

aridity, and an analysis of 224 dryland sites (Delgado-Baquerizo

et al

., 2013) highlighted a

negative effect of aridity on the concentration of soil organic C and total N, but a positive

effect on the concentration of inorganic P, possibly indicating the dominance of physical

processes such as rock weathering, a major source of P to ecosystems, over biological

processes that provide more C and N through litter decomposition (Delgado-Baquerizo

et al

2013).

Soil carbon dynamics in pastures strongly depend on management, with soil carbon increases

or decreases observed for different combinations of animal densities and grazing frequency

(Conant 2012; Machmuller

et al.

2015). Different grazing strategies, especially in the semi-

natural dryland biomes, have large implications for vegetation and the carbon balance (Yates

et al.

2000). Under certain conditions, grazing can lead to increased annual net primary

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production over un-grazed areas, particularly with moderate grazing in areas with a long

evolutionary history of grazing and low primary production but this does not always lead to

an increase in soil carbon (e.g. Badini

et al.

2007); grazing, like crop harvest, fundamentally

leads to the rapid oxidation of carbon that would otherwise be eventually transferred to the

soil. It has long been recognised that the potential effects of management on carbon storage in

grassland and dryland soils are substantially greater than that of climate change or CO

enhancement (Ojima

et al.

1993), and Henderson

et al.

(2015) estimated that the optimization

of grazing pressure could sequester 148 Tg CO

-1

4.4 Artificial surfaces, urbanisation and soil sealing

In 2014, 54% of the world’s population was urban, and by 2050, two thirds of the global

population will be urban. Many regions in the world, (such as Europe and Asia) are affected

by migration of populations from rural area to large megacities. Africa and Asia have more

rural populations, but are urbanizing faster than the other regions (World Urbanization

Prospects, 2014). With urbanization comes land take (development of scattered settlements in

rural areas, the expansion of urban areas around an urban nucleus, and densification on land

within an urban area) and soil sealing. Soil sealing refers to the permanent covering of an

area of land and its soil by impermeable artificial material (e.g. asphalt and concrete), for

example through buildings and roads. The area actually sealed is only part of a settlement

area, and gardens, urban parks and other green spaces are not covered by an impervious

surface (Prokop

et al

., 2011).

Sealing by its nature has a major effect on soil, diminishing many of its benefits (Tóth

et al

2007). It is normal practice to remove the upper layer of topsoil, which delivers most of the

soil-related ecosystem services, and to develop a strong foundation in the subsoil and/or

underlying rock to support the building or infrastructure. Loss of ecosystem and social

services (mainly on high-quality soils) include impacts on water resources (e.g. reduction of

rainfall absorbed by the soil, reduction of soil water holding capacity affecting flooding), on

soil biodiversity when sealing prevents recycling of dead organic material (Marfenina

et al.

2008), on the carbon cycle due to topsoil and vegetation removal (Davies

et al

., 2011).

Sealing through expansion of urban areas can also lead to agricultural land becoming more

marginal since the best agricultural land around settlements is lost as they expand, with

agricultural land displaced to more marginal land.

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Appropriate mitigation measures can be taken in order to maintain some of the soil functions.

In urban planning management, objectives to reduce the impact of soil sealing include: i)

preventing the conversion of green areas, ii) re-use of already built-up areas (e.g. brownfield

sites Meuser, 2010; Hester & Harrison, 2001 – though remediation of contaminated sites can

be costly; Maderova & Paton, 2013), iii) using (where appropriate) permeable cover materials

instead of concrete or asphalt supporting green infrastructure, and iv) implementation of

compensation measures. In order to deliver this mitigation a number of actions are necessary,

e.g. reduction of subsidies that act as drivers for unsustainable land take and soil sealing

(Prokop

et al

., 2011), and strong collaboration between relevant public authorities and

governance entities (Siebielec

et al.

2010). Development impacts can be reduced by

inclusion of green infrastructure, a network of high-quality green spaces and other

environmental features that have a positive effect on well-being (Gill

et al.,

2007) as well as

soils. In some regions, urban sprawl is exacerbated by insufficient incentives to re-use

brownfield (derelict, underused or abandoned former industrial or commercial) sites, putting

increasing pressure on greenfield land take.

Actions to alleviate pressures on soils driven by sealing fall into three categories: limiting,

mitigating and compensating. Actions to limit soil sealing centre around reduction of land

take through development of spatial urban planning and environmental protection. Mitigation

of soil sealing entails use of strategic environmental assessment for plans and programmes,

use of permeable materials and surfaces, green infrastructure within built and urban

environments, and natural water harvesting. Compensating soil sealing entails reclamation of

degraded land, re-use of extracted topsoil, de-sealing and is incentivised by land take fees and

application of environmental cost calculations.

Anthropogenic environmental change pressures that interact with land

management pressures on soils

In addition to the direct impacts of humans on soils

via

land use change and land

management, anthropogenic activity has indirect impacts through human-induced

environmental change, such as pollution and climate change. These interact with land

management. Soils provide a temporary but labile store for pollutants (Meuser, 2010).

Natural processes can release pollutants back to the atmosphere, make them available to be

taken up by plants and organisms, leached in to surface waters (Galloway

et al.

, 2008) and/or

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transported to other areas by soil erosion (Ravi

et al.

, 2010). Pollutants disrupt natural

biogeochemical cycles by altering both soil quality and function through direct changes to the

nutrient status, acidity and bioavailability of toxic substances and also by indirect changes to

soil biodiversity, plant uptake and litter inputs (EEA, 2014). Soil sensitivity to atmospheric

pollution varies with respect to key properties influenced by geology (cation exchange

capacity, soil base saturation, aluminium), organic matter, carbon to nitrogen ratio (C:N) and

water table elevation (EEA, 2014).

Atmospheric pollutant deposition impacts on soils vary with respect to soil sensitivity to a

specific pollutant and the actual pollutant load. Sulphur, nitrogen and heavy metals are

released in to the atmosphere by fossil fuel combustion (e.g. power generation, industry and

transport) and non-combustion processes (e.g. agricultural fertilizers, waste). These pollutants

are transported off-site and deposited as either dry or wet deposition, which can cross

national borders. Deposition is enhanced in forests and with altitude because of reduced wind

speeds and greater precipitation, respectively, impacting remote areas. Harmful effects to soil

function and structure occur where deposition exceeds the ‘critical load’ that a particular soil

can buffer (Nilsson & Grennfelt, 1988). Spatial differences in soil sensitivity (commonly

defined by the ‘crucial load’) and pollutant deposition result in an uneven global distribution

of impacted soils (Figure 3). For instance, global emissions of sulphur and nitrogen have

increased 3-10 fold since the pre-industrial period (van Aardenne

et al.

, 2001), yet only 7-

17% of the global land area sensitive to acidification is in a region where deposition exceeds

the critical load (Bouwman

et al.

, 2002).

Emissions of pollutants, notably sulphur, across Europe and North America have declined

since the 1980s following protocols established under the 1979 Convention on Long-range

Transboundary Air Pollution (LRTAP) and the 1990 US Clean Air Act Amendments

(CAAA) (Greaver

et al.

, 2012; Reis

et al.

, 2012; EEA, 2014). Conversely, emissions are

likely to increase in response to industrial and agricultural development in south and east

Asia, sub-Saharan Africa and South America (Kuylenstierna

et al.

, 2001; Dentener

et al.

2006). Further emission increases are occurring in remote areas due to mining activity, such

as oil sand extraction in Canada (Kelly

et al.

, 2010; Whitfield

et al.

, 2010).

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5.1 Sulphur deposition

Sulphur emissions are primarily from combustion of coal and oil, typically associated with

power generation and heavy industry. In 2001, regions with deposition in excess of 20 kg S

-1

where China and Republic of Korea, western Europe and eastern North America

(Vet

et al

., 2014; Figure 3a). Deposition in un-impacted areas is <1 kg S ha

-1

(Figure 3a).

Pollution control measures have seen an 80% reduction in pollutant sulphur deposition across

Europe between 1990 and 2010 (Reis

et al.

, 2012), and emissions in China have declined

since 2005 (Fang

et al.

, 2013).

Soil acidification is a natural process that is altered and accelerated by sulphur and nitrogen

deposition (Greaver

et al.

, 2012). Sulphur oxides (SO

) react with water to form sulphuric

acid (H

). Excess inputs of acidity (H

) displace soil base cations (e.g. calcium (Ca

) and

magnesium (Mg

)) from soil surfaces into solution, which are subsequently lost by leaching

(Reuss & Johnson, 1986). Mineral soils can buffer base cation losses if inputs from rock

weathering and/or atmospheric dust deposition exceed the amount lost. Therefore, the global

distribution of acid sensitive soils is associated with conditions that favour development of

soils with low cation exchange capacity and base saturation (Bouwman

et al.

, 2002; Figure

3c). Wetland can also buffer inputs of acidity through biological sulphate reduction, although

acidity can be mobilised again following drought and drainage (Tipping

et al.

, 2003; Laudon

et al.

, 2004; Daniels

et al.

, 2008). Organic acids can also buffer mineral acidity in naturally

acidic organic soils (Krug and Frink, 1983).

Decreased soil fertility or ‘sterilisation’ due to loss of nutrients and mobilisation of toxic

metals, particularly Al, are caused by acidification. Impacts in Scandinavia over the 1960s-

80s included declines in freshwater fish populations and damage to forests (EEA, 2014).

Sulphur can also stimulate microbial processes that make mercury bioavailable, leading to

bioaccumulation in the food chain (Greaver

et al.

, 2012). In agricultural soils in Europe,

however, fertilizer inputs of sulphur have increased to combat crop sulphur deficiencies as a

result of sulphur emission controls (Bender & Weigel, 2011).

Acidification is reversible, evident by increases in soil pH following decreased sulphur

emissions, although the recovery time varies; some areas with organic soils where deposition

has declined are showing either slow or no recovery (Greaver

et al.

, 2012; Lawrence

et al.

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2012; RoTAP, 2012). On agricultural soils, lime can be applied to increase soil pH.

However, 50-80% of sulphur deposition on land is on natural, non-agricultural land

(Dentener

et al.

, 2006). Application of lime to naturally acidic forest soils can cause further

acidification of deep soil layers whilst increasing decomposition in surface litter, with no

improvement in tree growth (Lundström

et al.

, 2003).

Wider effects of acidification are starting to be understood through long-term monitoring.

Decreased organic matter decomposition due to acidification has increased soil carbon

storage in tropical forests (Lu

et al.

, 2014). However, in temperate forest soils acidification

can lead to reduced C:N ratios in soil, which in turn increases nitrification (Aber

et al

., 2003),

but on already acidic soils reduces nitrification. In wetland soils, methane (CH

) emissions

have also been suppressed by sulphur deposition (Gauci

et al.

, 2004). Conversely, declining

sulphur deposition has been associated with increased dissolved organic carbon fluxes from

organic soils (Monteith

et al.

, 2007), and decreased soil carbon stocks in temperate forest

soils (Oulehle

et al.

, 2011; Lawrence

et al.

, 2012).

5.2 Nitrogen deposition

Nitrogen deposition covers a wider geographical area than sulphur, as the sources are more

varied, including extensive agriculture fertilizer application, ammonia derived from livestock

operations, and biomass burning in addition to fossil fuel combustion (Figure 3b). Regions

with deposition in excess of 20 kg N ha

-1

in 2001 were western Europe, South Asia

(Pakistan, India, Bangladesh) and eastern China (Vet

et al.

, 2014); although extensive areas

with 4 kg N ha

-1

were found across North, Central and South America, Europe and Sub-

Saharan Africa. By contrast, ‘natural’ deposition in un-impacted areas is around 0.5 kg N ha

-1

(Dentener

et al.

, 2006). While emissions related to fossil fuel combustion have declined

along with sulphur across Europe, agricultural sources of nitrogen are likely to stay constant

in the near future across Europe (EEA, 2014), whilst overall global emissions are likely to

increase (Galloway

et al.

, 2008). Nitrogen deposition in China’s industrialized and

intensively managed agricultural areas in the 2000s was similar to peaks in Western Europe

during the 1980s before mitigation (Liu

et al.

, 2013).

Deposition of nitrogen induces a ‘cascade’ of environmental problems, including both

acidification and eutrophication that can have both positive and negative effects on ecosystem

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services (Galloway

et al.

, 2003). Excluding agricultural areas where nitrogen is beneficial,

11% of land surface received nitrogen deposition above 10 kg N ha

-1

(Dentener

et al.

2006; Bouwman

et al

. 2002; Figure 3d). In Europe, eutrophication has and will continue to

impact a larger area than acidification (EEA, 2014).

Nitrogen fertilisation can increase tree growth (Magnani

et al.

, 2007) and cause changes in

plant species and diversity (Bobbink

et al.

, 2010), which in turn will alter the amount and

quality of litter inputs in to soils, notably the C:N ratio and soil-root interactions (RoTAP,

2012). However, increased carbon sequestration (Reay

et al

., 2008) may be offset by

increased emissions of the greenhouse gases N

O and CH

(Liu & Greaver, 2009). Long-term

changes caused by nitrogen deposition are uncertain as transport times vary between

environmental systems; and the only way to remove excess nitrogen is to convert it to an

unreactive gas (Galloway

et al.

, 2008).

5.3 Heavy metal deposition

Heavy metal emissions are associated with coal combustion and heavy industry. In the UK,

deposition is responsible for 25-85% of inputs to UK soils (Nicholson

et al.

, 2003). In

Europe, the areas at risk from cadmium, mercury and lead deposition in 2000 were 0.34%,

77% and 42% respectively, although emissions are declining (Hettelingh

et al.

, 2006).

Tighter legislation to control industrial emissions of heavy metals are helping to reduce the

environmental load of heavy metals in many regions, though rapid industrial growth in some

regions such as East Asia is increasing pressures on soil from heavy metal deposition. Global

heavy metal emissions and deposition are poorly understood in comparison to sulphur and

nitrogen; although the on-site impact of heavy metal contamination has been well studied

(Guo

et al.

, 2014). Metals in bioavailable form have toxic effects on soil organisms and

plants, influencing the quality and quantity of plant inputs to soils, rate of decomposition and,

importantly, can bio-accumulate in the food chain. Some heavy metals will persist for

centuries as they are strongly bound to soil organic matter (RoTAP, 2012), although they can

be mobilised to bioavailable form following drought-induced acidification, drainage and soil

erosion (Tipping

et al.

, 2003; Rothwell

et al.

, 2005).

Whilst the direct impacts of sulphur, nitrogen and heavy metals on inorganic soil chemical

processes are generally well understood, many uncertainties still exist about pollutant impacts

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on biogeochemical cycling, particularly interactions between organic matter, plants and

organisms in natural and semi-natural systems (Greaver

et al.

, 2012). Process understanding

is dominated by research in Europe and North America (e.g. Bobbink

et al.

, 2010). Research

is needed across Asia, Africa and South and Central America where soil properties and

environmental conditions differ. Models need to be developed to examine the combined

effects of air pollutants and their interactions with climate change impacts and feedbacks on

greenhouse gas balances and carbon storage (Spranger

et al.

, 2008; RoTAP, 2012). Air

quality, biodiversity and climate change polices all impact on soils. A more holistic approach

to protecting the environment is needed, particularly as some climate change policies (e.g.

biomass burning, carbon capture and storage) have potential to impact air quality and,

therefore, soil quality (Reis

et al.

, 2012; RoTAP, 2012; Aherne & Posch, 2013).

Indirect impacts on soils can be addressed largely by preventing the pollution at source, or by

mitigating the adverse effects where these have already occurred. Air pollution control on

coal burning and increased car and fleet efficiency standards has been effective in reducing

sulphur deposition in many areas of the world, particularly in Europe since the 1970s.

Substitution of coal with bioenergy might also reduce sulphur emissions, but unless burned

cleanly in a controlled way, can also release pollutants to the air. In terms of nitrogen,

ammonia abatement techniques when fertilizers are spread (e.g. slurry injection) are helping

to reduce N deposition (Sutton

et al

., 2007).

Existing policies and practices that alleviate global change pressures on soils

from land use and management

The previous text has highlighted specific anthropogenic activities that exert or alleviate

pressures on soils. Actions that alleviate pressures on soils driven by land use change and

land management can be broadly divided into three categories, those that:

1) Prevent conversion of natural ecosystems to other uses (e.g. protected areas, reduced

deforestation, prevention of wetland drainage, intensification rather than extensification);

2) Prevent soil degradation (erosion control, fire management, reduced tillage / conservation

agriculture, long term fallows, flood protection, use of organic amendments, intercropping,

improved rotations); and

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3) Result in soil / ecosystem restoration (e.g. peatland rewetting, afforestation, re-vegetation

on degraded lands, improved grass varieties, appropriate animal stocking densities,

bioremediation).

Policies to encourage such actions were recently reviewed by Bustamante

et al

. (2014) and

include:

a) Economic incentives, e.g., special credit lines for low carbon agriculture and forestry

practices and projects, payment for ecosystem services (such as carbon storage) and tradable

credits such as carbon,

b) Regulatory approaches, e.g. enforcement of environmental law to protect natural areas, set-

aside policies,

c) Research, development and diffusion investments, e.g. increase of resource use-efficiency,

livestock improvement,

d) Information and certification schemes, e.g. in China, forest certification to promote

sustainable forest management, state regulation for protecting mandatory arable lands,

protection projects on Tibetan grasslands, a national wetland protection programme, and the

“grain for green” programme.

Many of these actions and policies are not directed at soil conservation, but nevertheless have

an effect on soil quality. Two of the main pieces of international policy that have served to

reduce pressures on soils, directly and indirectly, are the United Nations Convention to

Combat Desertification (CCD) and the United Nations Framework Convention on Climate

Change (UNFCCC). In general, policies and actions are important at all scales from

international conventions to local action, and local activity is encouraged by policies at

regional, national and global level. Policies to sustainably increase land productivity, for

example, can prevent land use change, and there are various other supporting actions that can

help deliver improvements, e.g. agricultural research, technology transfer, knowledge transfer

and improved rural infrastructure. Some examples of policies that impact on land

management and soil quality are given below.

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6.1 United Nations Framework Convention on Climate Change (UNFCCC) and other

climate specific policies

Soil carbon storage and nutrient cycling as climate services are being increasingly recognised

e.g. under UNFCCC as part of national reporting and accounting, as part of life-cycle

greenhouse gas assessments for biofuels, in various regional initiatives and national efforts.

The UNFCCC is an international treaty, which came into force in 1994, setting an overall

framework for intergovernmental efforts to tackle the challenge posed by climate change.

The requirements for the 196 country Signatories (or ‘Parties’) to the UNFCCC include

adopting national mitigation policies and publishing national inventories of anthropogenic

emissions and sinks of greenhouse gases including activities on the land such as afforestation,

deforestation, agricultural management and wetland drainage and rewetting. Developed

country signatories have legally binding targets under the Kyoto Protocol and can count land

based emissions or sinks towards meeting these targets, thus incentivising activities that

protect soil carbon. Developing countries currently have voluntary targets and several

countries have made pledges that include reduced deforestation (e.g. Brazil and Indonesia) or

afforestation (e.g. 400000 km

in China). Under the Clean Development Mechanism (CDM)

developed countries can fund projects in developing countries that generate certified emission

reduction credits (CERCs). China, for example, has the largest number of CERCs in the

world (IFPRI, 2011). Brazil also has 180 CDM projects, the third largest number of CERCs

after China and India (Cole & Liverman, 2011). A number of projects in Africa, North

America and South Asia have a significant component for soil greenhouse gas emission

reduction of soil carbon sequestration, financed through the Verified Carbon Standard or the

American Carbon Registry.

As part of negotiations leading to the new climate treaty in Paris in December 2015, all

parties will be required to submit INDCs (Intended Nationally determined Contributions).

The new treaty will also include provision for REDD+ (reduced Emissions from

Deforestation and Degradation, including management of forests and enhancement of forest

carbon stocks). This could go some way to protecting forest soils, and negotiations have

been intense around methods for monitoring reporting and verification, with key issues such

as permanence (the risk the forest may be lost at a later date due to management or

environmental change) and leakage (displacement of land use change to other areas), and

how to finance such activities.

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6.2 United Nations Convention to Combat Desertification (CCD)

The CCD entered into force in December 1996; today 179 countries acknowledge it as a

legally binding framework to tackle land degradation and promote sustainable development

in fragile ecosystems. The Global Mechanism was established under the convention to

"promote actions leading to the mobilization and channelling of substantial financial

resources, including for the transfer of technology, on a grant basis, and/or on concessional or

other terms, to affected developing country Parties". In September 2011 the United Nations

General Assembly declared a goal of building a world with no land degradation. In October

2011 parties to the CCD issued a declaration calling for zero land degradation and for

adopting sustainable land management as a way to achieve sustainable development.

6.3 Millennium Development Goals (MDGs)

Of the eight MDGs (UNDP, 2014a), soil protection is most relevant to the goal to ensure

environmental sustainability, since soils are critical in underpinning environmental

sustainability (Smith

et al.

, 2015). A complementary MDG, to develop a global partnership

for development, will improve the governance structure to deliver soil security. The other

MDG to which soils plays a critical contribution is the goal to eradicate extreme poverty and

hunger, with the role of soils in supporting food provision critical for the latter part of this

MDG (Smith et al., 2015). The MDGs are currently being revisited to set a post-2015

development agenda (UNDP, 2014b), with discussion around the themes of localising the

post-2015 development agenda, helping to strengthen capacities and build effective

institutions, participatory monitoring for accountability, partnerships with civil society,

engaging with the private sector, and culture and development. The key emerging principles

from the dialogue are participation, inclusion, and the need for strengthened capacities and

partnerships (UNDP, 2014b). It is important that soils play their role in delivering this post-

2015 agenda.

6.4 Protected areas and the Convention on Biological Diversity (CBD)

Many measures to protect biodiversity and vulnerable habitats also protect the soils

underpinning them, so numerous conservation actions around the world serve to protect soils,

even if this was not the primary aim (Smith

et al

., 2013). Between 1990 and 2010, the

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amount of forest land designated primarily for the conservation of biological diversity

increased by 35 percent, indicating a political commitment to conserve forests. These forests

now account for 12 percent of the world’s forests (FAO, 2010). In India, a Supreme Court

ruling in 2011 on effective self-governance of “common” or communal land by local

communities may help to protect these valuable resources, and thereby the soils that underpin

them. Soil biodiversity is known to be important for soil function (Bodelier, 2011), yet it

rarely receives the attention enjoyed by larger flora and fauna within the ecosystem.

6.5 Reduced deforestation and forest management

Various actions have been implemented to reduce deforestation (Bustamante

et al

., 2014),

and to reduce the impact of forestry activities, such as reduced impact logging. UNFCCC,

carbon markets and other international environmental programs have contributed to global

efforts to reduce deforestation in addition to other sustainable natural resource management

programs in countries and by industry.

For example,

zero deforestation commitments made

by several companies (many made in 2014), and activities from bodies such as the

Roundtable for Sustainable Palm Oil (RSPO) and the Forest Stewardship Council (FCO)

certification scheme. Land improvement has increased in East Asia between 1981 and 2006

despite population increase, attributed largely to policies promoting tree planting and forest

plantation programs in China and Korea. In Brazil, deforestation was rapidly reduced after

national laws and regulations were enacted to protect forests in the 1990s and early 2000s

(including the soy moratorium and the forest code), followed up by state and municipal

governments setting further by-laws enforcing the deforestation moratorium (Bustamante

al.

, 2014).

6.6 Agricultural policies and practices

The pressures on soils imposed by land use intensity change can be mitigated by regulation of

over-grazing and reduction of over-stocking on grazed grasslands, return of crop residues to

the soil, reduced tillage, best management practices, targeted nutrient management and

precision farming on croplands, and wetland / floodplain restoration. These actions have been

encouraged by various policies. Some examples include: The EU set-aside programme of the

1990s encouraged less intensive use of agricultural land where production is low and

environmental impacts are high. The EU Common Agricultural Policy ties agricultural

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subsidies to implementation of best management practices and environmental protection, for

example through pillar 2 (rural development programmes) providing crop insurance for lower

fertilizer application rates; in Africa, policies for integrated land management to help protect

vulnerable soils; China’s conservation tillage program (2012-2030); the USA Conservation

Reserve Program (set aside marginal lands, steep slopes).

Conclusion: Keeping soils central to the science and policy agendas

The International Year of Soils in 2015 is an excellent opportunity to raise the profile of soils

in the minds of national and international policy makers, land managers, timber and agro-

industries, and the public. Ensuring that vulnerable and high environmental value soils (e.g.

peatlands) are considered when making policies and decisions about which habitats and

ecosystems to convert or to protect, will help to reduce the pressure on soils particularly

vulnerable to global change drivers such as land use and land management, and maintain

important ecosystem services. This is in part happening with agendas around valuation of

ecosystem services and life-cycle assessments of impacts of land use change that include soil

carbon. At a time when governments are negotiating a legally binding climate change treaty

and making national targets for greenhouse gas reduction, and revisiting the Millennium

Development Goals, keeping soil carbon and nitrogen central to land based greenhouse gas

monitoring and reporting will maintain awareness with policy makers and industries with

emissions reduction targets. Both science and policy agendas are increasingly concerned with

long-term food security, ensuring that soils are central to considerations of how to achieve

on-going increases in production will enable those increases to be more sustainable into the

future.

Research and policy regarding soil quality and sustainability is abundant, but patchy and

disjointed. To ensure that soils are protected as part of on-going wider environmental and

sustainable production efforts, soils cannot, and should not, be considered in isolation of the

ecosystems that they underpin, but the role of soils in supporting ecosystems and natural

capital needs greater recognition (Robinson

et al

., 2013, 2014). This can, in part, be enhanced

through education and awareness-raising which has started during the International Year of

the Soils in 2015. The time is ripe to consider a global soil resilience programme, under the

auspices of a global body such as the UN or one of its delivery agencies such as the FAO to

monitor, recover or sustain soil fertility and function, and to enhance the ecosystem services

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provided by soils. The lasting legacy of the International Year of Soils in 2015 should be to

bring together robust scientific knowledge on the role of soils, and to put soils at the centre of

policy supporting environmental protection and sustainable development.

Acknowledgements

The input of PS and PCW contributes to the Belmont Forum/FACCE-JPI funded DEVIL

project (NE/M021327/1) and for PS also contributes to the EU FP7 SmartSoil project

(Project number: 289694). The contribution of JIH was funded by the Leverhulme

Foundation. TAMP acknowledges funding from European Commission's 7th Framework

Programme, under Grant Agreement numbers 282672 (EMBRACE) and 603542 (LUC4C).

AKJ was supported by NSF (AGS 12-43071), DOE (DE-SC0006706), and NASA

(NNX14AD94G).

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and soil productivity response to corn residue removal: A literature review.

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biochar to mitigate global climate change.

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10.1038/ncomms1053.

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World

Bank, Washington, DC.

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Article

Zhang, X.H., Li, D.Y., Pan, G., Li, L.Q., Lin, F., Xu, X.W. (2008) Conservation of wetland
soil c stock and climate change of China. Adv.

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Tables

Table 1.

Observed and modelled soil carbon change (%) when converting from land cover

classes in the left hand column to land cover classes listed across the top. Results are from
meta-analysis of observations from the sources listed below. Model results (range across
three models) are shown for comparison in square brackets, range across the ISAM, LPJml,
and LPJ_GUESS models (see text), although note this calculated as difference in soil carbon
under the different land classes in 2010 and is thus not modelled loss/gain after a conversion.
Negative numbers represent loss of soil carbon.

Regrowth

Forest

Tree plantation

Grassland

Pasture Cropland

Forest

Global
Trop.

Temp.

Boreal

-9% (2)

-13% (3)

+8% (3)
-12% (2)

[-40 to -63%]

[-52% to +17]

[-14 to -49%]

-42% (3)
-41% (1)
-25% (2)

-30% (2)

-24% (5)
[-51 to -62%]
-52% (1)
-36% (4)
[-24 to -60%]
-31% (1)
[-63 to -65%]

Grassland Global

Trop
Temp

Boreal

[-1 to +15%]

[-28 to +3%]
[-26 to -71%]

[-2 to -6%]
-32% (4)
[-15 to -53%]
[-70 to -79%]

Pasture Global

Trop
Temp
Boreal

-10%

(3)

-59%

(3)

[-19 to +0.5%]
[-17 to -35%]
[-28 to -59%]

Cropland Global

Trop
Temp

Boreal

+53% (3)

+16% (4)

+18% (3)
+29% (2)
+20% (6)

+28% (4)

+19% (3)
+26% (2)

Footnotes:

Broadleaf tree plantations onto prior native forest or pasture did not affect soil C

stocks whereas pine plantations reduced soil C stocks by -12 to -15%;

Annual crops;

Perennial crops; 1 Wei

et al.

(2014a); 2 Don

et al.

(2011); 3 Guo & Gifford (2002; tropical

and temperate zones compiled); 4 Poeplau

et al.

(2011); 5 Murty

et al.

(2014); 6 Barcena

al.

(2014).

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Article

Table 2.

Soil carbon loss due to land use change 1860 to 2010 (PgCO

)

Model

Tropical

Temperate

Boreal

Global

LPJ-GUESS

46 55 1 109

LPJmL

128 95 0 227

ISAM

63 139 19 221

Mean

186

Table 3.

Threats to soil resource quality and functioning under increasing intensity of

agricultural management

Agricultural

management

practice

Specific

issue

Distribution Major

environmental

consequence

Knowledge

gap

Cropping

practice

Harvest

frequency

Global

Soil quality and

resilience

Impact on

total C and

nutrient cycles

Monoculture Global

but

particularly in

developing and

transition

countries

Soil health, pesticide

residue in intensively

managed

monocultures

Biological

resilience

Use of

agrochemicals

Over

fertilization

Particularly in

some developing

countries

Soil acidification,

water pollution, N

emission and nitrate

accumulation

Rate reducing

versus

balancing

Irrigation

Submerged

Rice

Developing

countries, Asian

Water scarcity,

methane emission

Trade-offs C

and water,

Arid/semi-

arid

regions

Arid/semi-arid

regions

Secondary

salinization, water

scarcity

Competition

use of water

Livestock

management

Over-grazing Largely

developing

countries

Soil degradation,

water storage, C loss

Forage versus

feed crops?

Industrial Largely

Waste pressure,

Safe waste

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Article

breeding industrialized

and

transition

countries

water pollution,

residue of veterinary

medicine and

antibiotics

treatment and

recycling

Agriculture in

wetlands

Wetland

drainage

Developing and

transition

countries

C loss

Agro-benefit

versus natural

value

Figure Legends

Figure 1.

Maps of change in soil carbon due to land use change land and land management

from 1860 to 2010 from three vegetation models. Pink indicates loss of soil carbon, blue
indicates carbon gain.

Figure 2.

Soil carbon and nitrogen under different land cover types in three different

vegetation models (values are the annual average over the period 2001 to 2010).

Figure 3.

Uneven global distribution of soils sensitive to pollution by (a) acidification and (b)

eutrophication (measured by soil C:N) compared to uneven distribution of atmospheric (c)
sulphur and (d) nitrogen pollution. Soils most sensitive to acidification have low base
saturation and cation exchange capacity, as defined by (Kuylenstierna

et al.

, 2001).

Acidification is caused by both sulphur and nitrogen. Eutrophication is caused by nitrogen.
Soil data in (a) and (b) were produced using the ISRIC-WISE derived soil properties (ver 1.2)
(Batjes, 2012) and the FAO Digital Soil Map of the World. Atmospheric deposition data in
(c) and (d) were provided by the World Data Centre for Precipitation Chemistry
(http://wdcpc.org, 2014) and are also available in Vet

et al.

(2014). Data show the ensemble-

mean values from the 21 global chemical transport models used by the Task Force on
Hemispheric Transport of Air Pollution (HTAP) (Dentener

et al.

, 2006). Total wet and dry

deposition values are presented for sulphur, oxidized and reduced nitrogen.

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