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Land use and land cover

Definition and transition

Land use (LU) refers to human activities such as the conversion of natural landscapes for human consumption or land management practices. This includes clearing forests, practicing intensive agriculture, shifting cultivation and expanding urbanization (Foley et al., 2005). LU activities occur to satisfy human demands for food, fiber, timber, shelter and other ecosystem goods.
The transition stages of LU from natural ecosystems to frontier clearings, then to subsistence agriculture and small-scale farms and finally, intensive agriculture and urbanization vary widely across the globe, as a function of their history, social, and economic conditions, and ecological characteristics (DeFries et al., 2004; Foley et al., 2005) (Fig 1.1). In particular, LU transition may occur from a system dominated by annual crops for local consumption to a system with large tree plantations in response to market demand (Lambin and Meyfroidt, 2010).

Forest transition

A forest transition is defined as a change in land cover trend, which is viewed as one of the important process in LU transition system (Lambin and Meyfroidt, 2010). Particularly, the concept of forest transition concerns the change of national forest areas from decreasing to expanding at a national or regional scale. The forest transition from deforestation to reforestation has varies in developed and developing countries (Lambin and Meyfroidt, 2010). According to the Global Forest


Assessment 2000, the world’s natural forest area decreased annually by 16.1 million hectares during the 1990s, which accounted for 4.2% of the total natural forest in 1990 (FAO, 2001). However, by the middle of the twentieth century, deforestation had essentially decreased in the world’s temperate forest as a consequence of the post-industrial economy based on the service sectors (FAO, 2012). Due to economic expansion, the labor force is driven from agriculture to other economic or service sectors and from rural to urban areas. The forest transition from decreasing to expanding forest density occurred mostly in Europe and in North America (Lambin and Meyfroidt, 2010). Nevertheless, the net global decrease in forest area was estimated at 9.4 million hectares per year over the decade from 1990 to 2000 (FAO, 2001). The forest transition of tropical regions also decreased. The FAO, 2010 estimated that 15.2 million hectares on average of tropical forest were lost per year during the 1990s. The FAO, in 2010, reported that the largest area of forest that was cleared in the world during 1990 – 2010 occurred almost exclusively in tropical regions (FAO, 2010) (Table 1.1).

Cropland expansion and agricultural intensification

Cropland expansion, the conversion from forest to cropland, is one of the most globally significant LU change (Foley et al., 2005). This is due to a combination of increasing food demand and market forces linked to global population growth. This increased production was made possible, in part, by the large scale clearance of major areas of forest: about 29% of forest and woodland was converted to agricultural uses from 1700 to 1992.
The rate of cropland expansion has slowed in the last three decades thank to agricultural intensification. Scientific and technological achievements have increased agricultural outputs through the use of high-yield crop varieties, chemical fertilizers, pesticides and herbicides and, irrigation systems (Matson et al., 1997). For example, tropical Asia increased its food production mainly by increasing fertilizer uses and irrigation. Agricultural intensification is expected to decrease the demand for cropland, thereby reducing the rate of conversion of natural ecosystems. This could lead to a stabilization of this LU transition or even an increase in tree cover.
In developed economies, a recent trend towards the conversion of agricultural land into non-agricultural uses has emerged. This transition towards urbanization, the abandonment of agricultural activity and the conversion of arable land or of permanent crops to other uses (pasture, woodland) is particularly evident in Western Europe or in the Northeastern United States where decreased in cropland have been observed (Lambin et al., 2003).


Urbanization also affects LU change through the transformation of the urban-rural linkage. The phenomenon of rural migration towards urban centers is driven by the perceived economic opportunities and by market forces (Satterthwaite et al., 2010). The rapid transition from rural to urban populations demands more resources and living spaces. Expanding cities are responsible for LU transitions such as forest clearing and the reduction of agricultural land surface to provide the surface needed for the establishment of the infrastructure required to support a growing population. Urbanization was in the past a prevalent phenomenon in developed countries (Glaeser, 2003). It is now almost complete; i.e. almost 80% of Europeans already live in urban areas (Haase et al., 2014). However, in developing economies urbanization continues to occur rapidly. Indeed, urban areas in developing economies are projected to increase from 300,000 km2 in 2000 to 770,000 km2 in 2030 and 1,200,000 km2 in 2050 (Angel et al., 2011).

Specifics of tropics and rural areas in developing countries

LU in rural tropical areas is rapidly changing from natural forest to perennial or annual crops (Ribolzi et al., 2016). High rates of tropical deforestation in developing economies are strongly linked to population growth and poverty. Growing population means increased demand on food productivity and increased demand for land for annual or permanent crops. Poor agricultural intensification and development increase pressure to convert forests and other marginal lands to crop production (Barbier, 2004). As a consequence, land expansions are expected to be higher in the agricultural sectors of developing economies in tropical regions.
In Asia, 65 % of forest loss was a result of LU change to agriculture (FAO, 2009) with 23% of this loss resulting directly from intensification of slash and burn agriculture, and 13% from direct LU conversions to small-size farms (FAO, 2009). In the early 20th century, about 90% of Southeast Asia was covered by forest and approximately 5% of the total land area was covered by cropland (Tao et al., 2013). From 1901 to 2000, the cropland area increased rapidly while forest area decreased. By 2005, forest area had decreased by 15.8% and cropland increased nearly 3 times when compared with the rate of increase in the early 20th century in the Southeast Asia. Most of these changes in LU occurred in Thailand, Indonesia, and the Philippines (Fig 1.2) (Tao et al., 2013).
Fertilizer application is one of the most popular agricultural practices in rural tropical countries to improve soil quality and productivity. The average rate of nitrogen fertilizer use in cropland increased from nearly 0 in the early 20th century to 2.8 g N m2 yr-1 in 2005 in tropical Asia (Tao et al., 2013). The use of chemical and organic fertilizers influences not only the soil and plant growth but also impacts the surrounding environment. Chemical fertilizer are cheaper, have higher nutrient contents and are rapidly taken up by plants (Robertson and Vitousek, 2009). However, many studies have indicated that the excessive use of chemical fertilizers has led to a range of environmental problems, particularly nutrient loss, soil acidification or basification, loss of useful soil microbial communities (Pernes-Debuyser and Tessier, 2004; Geisseler and Scow, 2014; Wang et al., 2017b) and surface water and groundwater contamination (Hallberg George, 1987; Easton and Petrovic, 2004).
The need for sustainability in agricultural practices has meant there is a movement toward the use of organic fertilizers as an alternative to mineral fertilizers. Organic amendments, including animal manure, solid wastes and various types of compost, biochar, are considered to improve soil quality and productivity (Glaser et al., 2014; Greenberg et al., 2017). Compared to chemical fertilizers, organic manure has lower nutrients contents and is often not optimized in terms of NPK concentrations and ratios (Gupta and Hussain, 2014). However organic fertilizers bring other benefits to soils due to maintenance of soil organic carbon, increasing soil microbial activity, improving soil structure and root development, and increasing soil water availability (Pernes-Debuyser and Tessier, 2004).

Consequences of land use change

LU activities impact biogeochemical cycles, cause biodiversity loss, modify and fragment habitats, degrade soil and water quality, and over-exploit native species. In streams, LU change causes more erratic hydrology and increased contaminant concentrations and affects ecological interactions (Lear et al., 2009).

Physical/ Hydrological perspective

One of the prominent consequences of shifting LU in tropical catchment systems is to increase soil erosion and overland flow that affects water quality in streams and rivers (Valentin et al., 2008). Rainfall – runoff is an important component of hydrological cycles. Overland flow (surface runoff) occurs when soils become saturated, which is common in low lying areas, or if precipitation exceeds the infiltration capacity of the soil (Kuchment, 2004). Raindrop impacts cause compaction of bare soil and increase soil particle detachment (Battany and Grismer, 2000; Mohamadi and Kavian, 2015). Surface roughness is one of the important factors affecting soil surface storage and infiltration, consequently controlling runoff and soil erosion processes during the rainfall (Ding and Huang, 2017). Differences in surface roughness can be due to soil texture and agricultural practices, as well as the deposition of residues or plant debris. Surface cover such as living vegetation or debris can reduce the impact energy of the raindrop, preventing the formation of surface crusts and enhancing infiltration capacity and hence reducing soil erosion (Battany and Grismer, 2000; Nunes et al., 2011).
Overland flow along with deep drainage (leaching) and subsurface drainage are the main pathways for the export of contaminants from land to water (Fig 1.3). The delivery of pollution to stream from hillslopes depends on the spatial and temporal movement of water via flow pathways and their connectivity (McKergow et al., 2007).
Fig 1.3: Water pathways from pasture to stream (left) and surface runoff generated by infiltration-excess and saturation-excess overland flow (right). Source: McKergow et al. (2007)
Soil erosion is a natural process involving the detachment and transport of soil particles by agents such as rainfall, runoff, wind, etc., (Lal, 2003). The soil erosion process is comprised of four stages: i) detachment of particles, ii) breakdown of aggregates, iii) transport and redistribution of sediments over the landscape and iv) deposition in aquatic ecosystems (Lal, 2003). The transfer of solute from the soil surface to overland flow includes the transfer of solutes from the soil surface by diffusion induced by the concentration gradient, ejection of solution from the soil surface by raindrops, erosion by raindrops and surface flow of sediment with adsorbed chemicals, and adsorption–desorption of the adsorbing chemicals (Fig 1.4) (Shi et al., 2011).
Overland flow, and hence the rate of soil erosion as well as solute transfer processes from soil surface to overland flow are controlled by local site characteristics, particularly soil texture, structure and properties, vegetation type and degree of cover, topography and rainfall (Kuchment, 2004; Montgomery, 2007; Shi et al., 2011). Infiltration rate is closely related to soil texture. For example, coarse textured soils have higher infiltration rates than fine texture soils due to large and well-connected pore spaces. Topographical factors, including slope gradient and slope length, directly affect the potential energy and the nature of overland flow. Runoff volume increases with the increase of slope gradient and length (Shi et al., 2011). Vegetation and the degree of surface crusting also contribute to determine overland flow rates with high vegetation coverage and low crusting leading to higher infiltration as compared to bare soils with high crusting rates (Kuchment, 2004). The links between LU change and agricultural practices, soil erosion and overland flow have been found in many studies in both temperate and tropical regions (Montgomery, 2007; Valentin et al., 2008; Nunes et al., 2011; Lacombe et al., 2017). For example, Nunes et al. (2011) found that the soil loss by erosion for the cereal crop was more than 20 times higher than that of fallow land or short-term abandoned land. The formation of shrub cover and recovering oak forests resulting from land abandonment showed low rates of runoff (1.05 mm) and soil erosion (0.62 g m-2) due to the high infiltration capacity compared to cereal crop with 77.7 mm of runoff and 627.1 g m-2 of soil erosion (Nunes et al., 2011).
Shifts towards different LU can cause changes in soil erosion, as well as loss of soil organic matter and nutrient concentrations (Navas et al., 2012). For example, the change of forest to maize crops caused the loss of soil organic carbon (SOC) by 20%, 36%, and 47% after three, seven, and ten years in eastern Thailand (Jaiarree et al., 2011). Niu et al. (2015) when evaluating soil erosion and variation in SOC, total nitrogen (TN) and total phosphorous (TP) contents under different LU types in the catchment of the Dianchi Lake of China found that grasslands and forestland experienced a lower degree of soil erosion and higher SOC, TN and TP compared to tillage or abandoned agricultural lands. Similarly, the conversion form native tropical forest (Kenya) and subtropical grassland (South Africa) ecosystems to agriculture led to the depletion of the total SOC (46-73%) and N concentration (37-73%) during the first 4 years of conversion (Solomon et al., 2007). Tiessen H (1992) reported that 30% of carbon, nitrogen, and phosphorous of soil were lost when applying slash and burn agriculture during 6 years in tropical countries. Other authors found that 8–10 years of fallow were needed to restore fertility levels to those similar to the original site conditions prior to cultivation (Matson et al., 1997). Similarly, the conversion from cultivated land to urban areas decreased organic production and nutrient cycling such as phosphorus and potassium due to the loss of vegetation (Song and Deng, 2015). Wang et al. (2017a) also indicated that LU change driven by urbanization influences the microbial processes involved in nitrogen cycling.
Soil erosion and overland flow processes that are accelerated by LU and agricultural practices are responsible for transporting sediments, carbon, nutrients and other chemical fluxes from soils to  surface waters (Kuchment, 2004). Hughes et al. (2012) found that mixed LU catchments export up to three times more sediment than the native forest catchments. Similarly, Huang and Lo (2015) indicated that the 6.9% decrease of forest land and 9.5% increase of agricultural land caused an increase in sediment yield of 0.25 t ha-1 over the whole river basin.
SOC is easily transported by runoff water due to its relative low density (<1.8 Mg/m) and because it is concentrated in the vicinity of the soil surface (Lal, 2003). Leaching losses of dissolved organic carbon (DOC) may also be an important pathway of carbon loss from agricultural systems (Kindler et al., 2011). The overall mean DOC exported from the pasture and mixed (pasture, pine, and native) catchment were 50% higher than the native forest catchment (Quinn and Stroud, 2002). This loss of organic carbon from soils through erosion and runoff can have serious consequences for agricultural lands, as well as on the downstream water bodies to which it is imported (Cole et al., 2006). For example, Walmsley et al. (2011) found that carbon leaching from soils reduced potential soil carbon gains by 20 g m-2 y-1. Additionally, higher DOC concentration impacted the transport of nutrients and pollutants such as enhancing the sorption and mobility of pesticides and heavy metals (Veum et al., 2009; Pagano et al., 2014). This leads to degrade the drinking water quality. Moreover, intensive agriculture such as the excessive application of fertilizers has also become the largest source of excess of nitrogen and phosphorous to aquatic ecosystem (Matson et al., 1997; Bennett et al., 2001; Foley et al., 2005).

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Sources, forms and characteristics of organic carbon

Sources and forms

Organic carbon in water can be classified into two categories: particulate (>0.7 µm) and dissolved (<0.7 µm) (Nachimuthu and Hulugalle, 2016). DOC is a general description of the organic carbon dissolved in water. Dissolved organic matter (DOM) is a term used generally for dissolved organic substances, but the term is often used interchangeably with DOC in the literature, despite the fact that DOC makes up a fraction of the DOM profile (Pagano et al., 2014). There are two sources of DOC in aquatic systems: allochthonous (external to the system) and autochthonous (internal to the system). Autochthonous DOC is released from primary production from macrophytes and benthic and phytoplankton (Thomas et al., 2009; Bauer et al., 2011) through different mechanisms including predator grazing, cell death, viral lysis and extracellular release (Azam et al., 1983). Meanwhile, allochthonous DOC is derived from terrestrial vegetation and soil. It is estimated that about 2.9 Pt C yr-1 of the terrestrial organic C enters freshwater systems worldwide (Tranvik et al., 2009). Terrestrial plants and soil organic matter are the major sources of carbon in temperate headwater streams (Vannote et al., 1980). The quantity and quality of allochthonous DOC that enters aquatic ecosystems is controlled by many factors such as the hydrologic conditions of the stream, LU management and agricultural practices, precipitation, soil type, vegetation cover, etc. (Willett et al., 2004). For example in New Mexico, USA, forest soils tend to lose more organic carbon than grassland soil, with on average 39 g and 5 g lost respectively (Puttock et al., 2013). Williams et al. (2010) in the survey of 43 streams distributed throughout catchments of mixed LU in Southern Ontario, Canada also indicated that DOM in agriculturally affected streams was likely more labile and accessible to the microbial community than DOM in the wetland and forest affected streams.

Table of contents :

1.1.􀀁 Land use and land cover
1.1.1.􀀁 Definition and transition
1.1.2.􀀁 Specifics of tropics and rural areas in developing countries
1.2.􀀁 Consequences of land use change
1.2.1.􀀁 Physical/ Hydrological perspective
1.2.2.􀀁 Sources, forms and characteristics of organic carbon
1.3.􀀁 Aquatic microbial community
1.3.1.􀀁 Free living (FL) and Particles Attached (PA) bacteria
1.3.2.􀀁 Factors controlling microbial community
1.3.3.􀀁 Microbial degradation of DOC
1.3.4.􀀁 Links between bacterial diversity and functioning
2.1.􀀁 Study sites
2.1.1.􀀁 Dong Cao catchment
2.1.2.􀀁 The Houay Pano catchment
2.2.􀀁 Field work
2.2.1.􀀁 Rain simulations and mesocosm/microcosm experiments
2.2.2.􀀁 Sampling
2.3.􀀁 Laboratory work
2.3.1.􀀁 Measurement of DOC, nutrients and Chlorophyll a
2.3.2.􀀁 Measurement of CDOM
2.3.3.􀀁 Measurement of metabolic capacity
2.3.4.􀀁 DNA extraction
2.4.􀀁 Sequence processing and data analysis
2.5.􀀁 Statistical analyses
3.1.􀀁 Introduction
3.2.􀀁 Methods
3.3.􀀁 Results 63􀀁
3.4.􀀁 Discussion
4.1.􀀁 Introduction
4.2.􀀁 Material and methods
4.3.􀀁 Results.
4.4.􀀁 Discussion
4.5.􀀁 Conclusion
5.1.􀀁 Vicinal land use change strongly drives stream bacterial community in a tropical montane catchment
5.1.1.􀀁 Introduction
5.1.2.􀀁 Materials and Methods
5.1.3.􀀁 Results.
5.1.4.􀀁 Discussion
5.1.5.􀀁 Conclusion and perspectives
5.2.􀀁 Impact of overland flow on particle-attached and free-living fractions during a flood event in an upland tropical catchment
5.2.1.􀀁 Introduction
5.2.2.􀀁 Material and methods
5.2.3.􀀁 Results.
5.2.4.􀀁 Discussion
5.2.5.􀀁 Conclusion
6.1.􀀁 Introduction
6.2.􀀁 Material and methods
6.3.􀀁 Results and discussion
6.4.􀀁 Conclusion


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