Phosphorus cycling in boreal and temperate forest ecosystems

Transcription

1 Phosphorus cycling in boreal and temperate forest ecosystems A review of current knowledge and the construction of a simple phosphorus model A report from Belyazid Consulting & Communication

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3 A report from Belyazid Consulting and Communication AB December 2012 Authors Ulrika Jönsson Belyazid & Salim Belyazid, Belyazid Consulting and Communication AB, Österportsgatan 5c, Malmö, Sweden Contact Ulrika Jönsson Belyazid, Tel: Acknowledgements This work was funded by the Swiss Federal Office for the Environment. We thank Sabine Braun, Institute for Applied Plant Biology, Switzerland, for useful comments on previous versions of the report and for providing data. The photos in the report are from dreamstime.com and freedigitalphotos.net (except for photos on pages 6 and 21). The photo on page 21 was kindly provided by Sabine Braun.

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5 SUMMARY Phosphorus (P) is of vital importance for both plant and animal life and is a critical element in natural and agricultural systems throughout the world. In order to manage P for economic plant production and environmental protection, we need to understand the nature of the different forms of P found in plants and soils and the manner in which these interact with various components of the ecosystem and the surrounding environment. This report identifies the processes controlling P dynamics in temperate and boreal forest ecosystems and investigates how these processes are influenced by various environmental conditions. The information used in the review comes mainly from articles that have been published in international scientific journals and both empirical and modelled data have been included. To simplify, the processes have been divided into P inputs, P in the soil, P in the plants and P outputs. The information gathered in the review was then used to first create a conceptual model and then a prototype model for P cycling in boreal and temperate forest ecosystems (the latter using empirical data). 1. P inputs P enters forest ecosystems through atmospheric deposition and weathering, and sometimes also in the form of fertilizer. The deposition of P is generally small (0,05-1,7 kg P ha -1 yr -1 ), but may nearly balance the losses from the soil in undisturbed forest ecosystems. With regard to P in throughfall, results are variable, with some studies showing an uptake of P in the tree crown while others report leaching of P from the crown. Whether there is an uptake or depletion have been related to tree species, season and the nutrient supply of the plants. Age of foliage and canopy closure has also been suggested to influence the outcome. A large fraction of the P occurs in the rock as apatite, and the rate of P release by chemical weathering has been estimated to be between 0,05 and 1,0 kg P ha -1 yr -1, with a possible range of up to 5 kg P ha -1 yr - 1. Weathering is generally increased by increasing soil temperature and moisture, and is also influenced by protons and chelating agents produced by plants and microorganisms. One factor that is likely to be important for determining how much of the P released by weathering that is able to circulate in the forest ecosystem is rooting depth. Fertilization with P is a common practice in many parts of the world. Generally, the P in the fertilizer is rapidly sorbed to soil particle surfaces, and only a limited amount of the P added is likely to be taken up by plants in the year of fertilization. 2. P in the soil The bulk of soil P exists in three general groups of compounds: organic P, calcium (Ca) bound inorganic P and iron (Fe) or aluminium (Al) bound inorganic P. The relative amounts of organic and inorganic P vary greatly from soil to soil, and suggestions on the fraction of organic P range from 5 to 95%. Both organic and inorganic P contribute P to the soil solution, but most of the P in each group is of very low solubility and not readily available for plant uptake. The concentrations in soil solution usually range from 0,001 mg l -1 in very infertile soils to about 1 mg l -1 in rich, heavily fertilized soils. There are a number of organic P compounds present in the soil, but their identity and amounts are not well understood. Inositol phosphates are currently believed to be the most abundant organic P compound in the soil. However, the suggestions on the fraction they constitute (10 to 80%) indicate the uncertainty of these estimates. P mineralization and immobilization are controlled by many of the same processes that control the general decomposition of soil organic matter, such as temperature, moisture and availability of nitrogen. The carbon/phosphorus ratios under which net mineralization has been suggested to occur are highly variable, ranging from 200:1 til 1200:1. Mineralization of organic P in soils in temperate regions typically release between 5 and 20 kg P ha -1 yr -1. The P that is mineralized enters the soil solution. There, it may be reimmobilized into organic matter, taken up by plants or adsorbed or fixed into insoluble forms by reaction with Fe, Al, Ca and clays. The particular types of reactions that fix soluble inorganic P into relatively unavailable forms differs from soil to soil but are generally closely related to ph. As a general rule of thumb, plant availability of P is regarded to be highest when soil ph is between 6 and 7. In most alkaline soils, Ca compounds dominate, while Al and Fe forms are more important in acid soils. In both acid and alkaline soils, P tends to undergo sequential reactions that produce P-containing compounds of lower and lower solubility. Most of the compounds with which P reacts are in the finer soil fractions. If soils with similar ph values and mineralogy are compared, P fixation thus tends to be more pronounced in soils with higher clay contents. 3. P in the plants Plants absorb P dissolved in the soil solution mainly as orthophosphate (preferentially H 2 PO 4 - ). There is hitherto no evidence for direct uptake of organic P compounds by plants and inorganic P is also the only form taken up by mycorrhizal fungi. As a response to the low availability of P in the soil, plants have developed a number of different strategies to increase their uptake efficiency of P. These mechanisms can roughly be divided into two groups: mechanisms that increase the plants uptake capacity of P (root architecture and microbial associations such as mycorrhiza) and mechanisms that change the equilibrium between different P compounds in the soil, thereby increasing the soil availability of P (exudation of protons, phosphatases and organic acids). Many of these mechanisms seem to operate on several levels and influence the availability and uptake of both organic and inorganic P. The responses seem to be controlled by both P status in the shoot and the availability of inorganic P in the rhizosphere. The concentration of P and the ratio between P and other nutrients in foliage of trees may vary substantially and is influenced by several different factors, among others tree age, N deposition and drought. Generally, critical concentrations and ratios are specifically related to growth. However, adequacy ranges for nutrient concentrations and ratios should also take other ecological aspects into account, since nutrient status may affect processes such as cold hardiness, drought resistance and susceptibility to parasite attacks. One of the most recent and comprehensive studies on P nutrition of trees has suggested optimum P concentrations between 1,2 and 2,2 mg P g -1 dry weight and N/P ratios between 6,3 and 19,6, depending on tree species and based on the growth responses. Eventually these values agree well with those taking into consideration also other abiotic and biotic site conditions. Many of the responses to P starvation appear to be initiated or modulated by a decrease in the delivery of inorganic P to the shoot and the consequent reduction in inorganic P available for shoot metabolism and shoot growth. In contrast to shoot growth, root growth is less inhibited under P deficiency and most species partition a greater proportion of their total dry matter into root growth when grown under P deficiency, typically resulting in an increase in the root to shoot ratio of the plant when P is scarce. However, eventually, photosynthesis is also impaired. P fertilization generally results in both improved nutritional status and increased growth of trees. The effects of P fertilization generally last for long (more than 20 years) and P fertilization has in some cases even been found to have a positive impact on growth in subsequent rotations. However, there are also studies showing no effect on growth despite improved P nutrition of trees as a consequence of fertilization. P resorption from leaves is an important conservation mechanism for P, although the rates seem to vary considerably between studies (0-80%). The resorption efficiencies do not seem to be clearly related to soil fertility. Instead, internal factors, such as growth rate, have been suggested to control the resorption efficiency. In coniferous forests, P input to litter (in g m -2 ) is dominated by P from needles. In deciduous forests, fruits may also contribute significantly to

6 the input of P. The contribution from coarse woody debris, on the other hand, is generally considerably smaller because of its low nutrient content. P fertilization has been shown to increase the P content of litter and increase the return of P by 150 to 400% compared to unfertilized forests. 4. P outputs The principal pathways by which P is lost from an ecosystem are erosion of P carrying soil particles, P dissolved in surface runoff water and plant removal. Under good vegetation cover (>60%), soil erosion is generally limited. However, disturbances such as timber harvest or wildfires may increase the loss of P substantially. Soils with medium to fine texture, low organic matter content and weak structural development is generally most easily eroded. Erosion of P-carrying soil particles may vary between 0,1 and 10 kg P ha -1 annually, the higher value most likely applies only to cultivated soils. Because soluble inorganic forms of P are strongly adsorbed by mineral surfaces, the loss of P through leaching is generally very low (<0,7 kg P ha -1 ). Hitherto, there are no well established controls of P leaching from forested areas. The major loss of P as a consequence of forest management is the direct removal of P in tree parts. Plant harvest may remove approximately 5 to 50 kg P ha -1 annually from an ecosystem, depending on plant species, soil type and cultivation system. Whole-tree harvest may remove the double, sometimes even more. Although indirect effects of forest management practices are negligible on a larger scale, they may have significant impact on P export on a local scale. 5. Prototype model of the P cycle The model prototype, that was built as a result of the review, focuses on four forms of P: P in the biomass, mineral P in the soil, organic P in the soil, and P in the soil solution. The model simulates the balance between the four P pools based on the sizes of the flows affecting them, which in turn depend on the strengths of the processes controlling the latter. The dissolution and precipitation of P under inorganic forms is simulated according to a Langmuir saturation function. Soil organic P is released through the decomposition process, and subject to retention by microbes. P in the plants is steered by the minimum nutrient requirements in the plant, and feedback through regulating photosynthesis and growth, particularly under deficient status. Finally, P in the soil solution is short lived, as it is readily taken up by microbes and plants. The model does not represent the leaching of particulate P, as we lack descriptive information on this process at present. 6. Conclusions and future perspectives The literature review provided sufficient information to synthesize a complete conceptual model for the P cycle in temperate and boreal forest ecosystems. However, it also showed that there is a considerable lack of empirical data numerically describing the flow and transformation rates of P, particularly in the soil. Nevertheless, the information presented in this report shows that it is possible to include the P cycle in existing biogeochemical models with reasonable accuracy. To account for the P cycle is vital to the reliability of existing biogeochemical models and the accuracy of their predictions, particularly considering the scarcity of P in natural and semi-natural ecosystems. Although the study summarized and synthesized a large amount of existing information, it points to the necessity of investigating the effects of environmental changes and human interference on the P cycle and long-term availability of P in non-agricultural ecosystems. From the perspective of managed boreal and temperate forest ecosystems, investigating the response of the P cycle to direct and indirect climate change effects and looking closer at the forest management practices topical to meet the challenge of climate change is of significant interest.

7 TABLE OF CONTENTS 1. Introduction 7 2. Aims and objectives 8 3. Report layout and method 9 4. P input Atmospheric deposition of P Throughfall 4.2 Weathering P fertilization P in the soil Forms and content of P in soils Mineralization and immobilization of P Effects of climate Effects of N Effects of nutrients other than N Effects of aluminium and ph Effects of parent material 5.3 Adsorption and desorption of P Effects of ph Effects of clay content Effects of organic matter Effects of age 6. P in the plants P acquisition and uptake Availability of P in the soil solution Forms of P taken up by plants Root uptake Leaf uptake 6.2 P content of plants and its relation to growth Different forms of P in plants P content in plants P concentrations and nutrient ratios P and plant growth Effects of P fertilization on P concentration and growth 6.3 Retranslocation of P P in litterfall P output P erosion Leaching of P Effects of tree harvest and forest management on P A conceptual model Causal loop diagram a tool to visualize conceptual models A CLD of P cycling in temperate and boreal forest ecosystems The prototype model Conclusions Future perspectives References 33

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9 1. INTRODUCTION Phosphorus (P) is a macronutrient that is a key component of several cellular compounds, among others adenosine triphosphate (ATP), deoxyribonucleic acid (DNA) and ribonucleic acid (RNA). ATP is a highenergy phosphate group that drives most energy-requiring biochemical processes within the plant, while DNA is the seat of genetic inheritance and RNA the director of protein synthesis. P is thus of vital importance for both plant and animal life and is a critical element in natural and agricultural ecosystems throughout the world. An adequate supply of P enhances many aspects of plant physiology, including the fundamental processes of photosynthesis, nitrogen (N) fixation, flowering, fruiting and root growth. Deficiency of P, on the other hand, often results in stunted, thin-stemmed spindly plants with a dark, almost bluish, green colour, although symptoms may vary substantially depending on the surrounding circumstances (Brady & Weil, 1999; Marschner, 2003). The natural supply of P in most soils is small, and the availability of that which is present is generally very low (Brady & Weil, 1999). Natural ecosystems have thus evolved in ways that promote certain biological and chemical processes that allow plants to make efficient use of the scant supplies available. However, human activities have greatly altered the biogeochemical cycles of Earth (Aber et al., 1989; Galloway, 1998), increasing the availability of N in many natural ecosystems. With regard to forests in boreal and temperate areas, several studies have suggested that a high availability of N as a consequence of high atmospheric deposition or fertilization may result in imbalances between N and P, or even deficiencies of P (Houdijk & Roelofs, 1993; Binkley & Högberg, 1997; Flückiger & Braun, 1998; 1999; Harrison et al., 1999; Duquesnay et al., 2000; Braun et al., 2010; Prietzel & Stetter, 2010; Blanes et al., 2012). During recent decades, a significant deterioration of the P nutrition and an increase in the N/P ratios of European forests has been noticed (Duquesnay et al., 2000; Jonard et al., 2009; Braun et al., 2010; Prietzel & Stetter, 2010), although there are exceptions (Finzi, 2009; Groffman & Fisk, 2011). According to Jonard et al. (2009), the changes commonly observed can be explained by the joint action of several processes that influence tree nutrition in the long term, namely tree age, N and sulphur (S) deposition, harvesting and climate. Unless sources of available P can be mobilised, or added, growth of the trees may become poor, with consequences for forest health, productivity and the ecosystem services usually provided by forests. At the other extreme, the liberal application of P over many decades in order to support human food production have resulted in increased levels of this element in the surface layers of many agricultural soils. Also, human, animal and industrial wastes tend to concentrate P. Over large parts of the world, P lost from watersheds and waste treatment facilities in surface runoff water and eroded sediments is severely upsetting the nutrient balance in streams, lakes and estuaries, the consequence being eutrophication. This may jeopardize both drinking-water supplies and can severely restrict the use of aquatic systems for fisheries, recreation and aesthetics (Brady & Weil, 1999). In order to manage P for economic plant production and environmental protection, we need to understand the nature of the different forms of P found in plants and soils, and the manner in which these interact with the components of the ecosystem and the surrounding environment. To be able to explain and foresee the development of P deficiencies in forest ecosystems, including the P cycle in dynamic forest ecosystem models becomes imperative. With the exception of agricultural soil models, P remains missing from most terrestrial ecosystem models, which have been traditionally focused on the cycles of carbon (C), water, N and, to a more limited extent, base cations. The present review is an attempt at collecting available empirical and modelled evidence describing the different segments of the P cycle in forests and the processes governing them. This information forms the basis of the model structure as well as its parameterisation. Runoff 0,3 0,18 0,12 Weathering Plant uptake and litter Precipitation An example of a P balance in a forested watershed, where the forest consisted primarily of mature hardwoods that had remained undisturbed for 45 years or more. The soil is the surface soil of an Ultisol. Flows of P, represented by arrows, are given in kg ha -1 yr -1. Modified from Brady & Weil (1999). 0,14 Organic P 645 kg/ha Inorganic P 275 kg/ha Mineralization Nitrogen (mg g -1 d.m.) Fagus Picea Phosphorus (mg g -1 d.m.) Fagus Picea N:P ratio (w/w) Fagus Picea Year Year Year Development of foliar concentrations of N (left), P (middle), and N/P ratio (right) of mature beech and Norway spruce in permanent forest observation plots in different regions of Switzerland between 1980 and Values are corrected for age trend. Bars = 95% confidence intervals of plot medians. Modified from Braun et al. (2010). 7

10 2. AIMS AND OBJECTIVES The aims of this report are: To identify what processes are controlling P dynamics in temperate and boreal forest ecosystems. To investigate how these processes are influenced by various environmental conditions. 8 To create a conceptual model for P cycling in temperate and boreal forest ecosystems. To create a prototype model for P cycling in temperate and boreal forest ecosystems, based on the conceptual model and using empirical data.

11 3. REPORT LAYOUT AND METHOD The report starts with a literature review of the major pathways and processes governing P in temperate and boreal forest ecosystems. The review is divided into four sections covering P inputs into ecosystems (section 4), P in plants (section 5), P in soils (section 6) and P outputs from ecosystems (section 7). The literature review is based on studies that have been published in international scientific journals. We have mainly included information of relevance with regard to temperate and boreal forest ecosystems in Europe. However, when the information available was scarce, we extended the literature search to other types of forest ecosystems and other regions of the world. We have also included some information specific to alpine ecosystems, where such data was available. Both empirical and modelled data have been included in the review. We are well aware of the difficulty in drawing coherence from a large number of studies, which to varying degrees, are site-, age- and species-specific. However, we have tried to include as much information as possible within the limited time available. Although very important (see for example Suárez et al., 2003 and Chapuis-Lardy et al., 2011), the role of the soil macrofauna in P cycling has been largely been ignored in this review. Hopefully, we will be able to take a closer look at the impact of soil macrofauna on forest nutrient cycling in future projects. Furthermore, no description or evaluation of soil test methods for determination of available P in soils is included, since the subject is complicated enough to warrant a review of its own. Based on the literature review, we have created a conceptual model for P cycling in temperate and boreal forest ecosystems (section 8). The model is presented in the form of a Causal Loop Diagram (CLD). The CLD shows how the different transformations of P take place in the ecosystem, and what variables control the different processes. CLDs are widely used in system analysis and system dynamics studies, where complex multidimensional problems require a holistic approach and studies cannot be carried out solely on isolated parts of the system (Haraldsson, 2004). The CLD was then used as a blueprint to create a prototype numerical model (section 9). The model was built in the STELLA modelling environment. In this STELLA prototype model, the variables and processes described by the CLD in Section 8 were translated into stocks, flows and processes describing the P cycle in a forest ecosystem, as we understand it at the issue of this study. The STELLA model is a prototype of the P cycle alone, and is the basis for including such a cycle in a fully integrated forest ecosystem model at a later stage. The qualitative dynamics displayed by the prototype are based on the findings of this study, while the quantitative numerical parameterisations are based on reviewed empirical data and to a large extent on data provided by Sabine Braun, Institute for Applied Plant Biology, Switzerland. The STELLA prototype allows us to test the hypothesis describing the processes governing the P cycle. It is also a tool to test the numerical description of the P cycle, i.e. how the sizes of the different P flows and stocks as well as the rates of the different reactions and processes interact to create a realistic P budget for a modelled forest stand. The STELLA prototype contains the equations that will eventually be included in a larger biogeochemical model where interactions with other element cycles will be considered. Some conclusions and future perspectives can be found in section 10 and references in section 11. 9

12 4. P INPUT 4.1 Atmospheric deposition of P The amount of P that enters the soil from the atmosphere is generally small (Attiwill & Adams, 1993; Newman, 1995; Brady & Weil, 1999). Of the P deposited from the atmosphere, some is dissolved in rain, mist or snow, while some is in particles. The particles may be suspended in rain, mist or snow, or may be separate dry particles (dry deposition). Some of the P is likely to be in organic forms, while some is in inorganic forms (Attiwill & Adams, 1993; Newman, 1995; Brady & Weil, 1999). There are several types of technical problems in measuring P inputs to terrestrial ecosystems (see for example discussion by Newman, 1995) and the numbers presented in the literature are thus highly variable. Newman (1995), reviewing many of the existing studies on P inputs to terrestrial ecosystems, concluded that inputs from the atmosphere range from 0,07 to 1,7 kg P ha -1 yr -1. However, he emphasized that aerosols impacting on foliage may increase these numbers substantially at some sites. Other sources have suggested P deposition to range from 0,05 to 0,5 kg P ha -1 yr -1 (Brady & Weil, 1999). Of this amount, dry deposition may constitute a significant part (e.g. 89% in one study reviewed by Newman, 1995). Since aerosol impaction is affected by the structure of the canopy, conifer forests would be expected to trap more aerosols than dicotyledonous forests and forests more than grasslands (Newman, 1995). For Switzerland, deposition rates of P to agricultural land has been estimated to 0,4 kg P ha -1 yr -1 (Spiess, 2011). Although relatively low, the amount added through deposition may nearly balance the losses from the soil in undisturbed forest and grassland ecosystems (Brady & Weil, 1999). Kopácek et al. (2011a) reported than the terrestrial P exports from catchments along an elevation gradient in the Tatra mountains and Bohemian forest in the Czech republic was generally lower than atmospheric input of P to the soils. In accordance with Kopácek et al. (2011a), Zhang & Mitchell (1995) found that leaching of total P from the soil in a hardwood stand in the eastern US was only 26% of the input by precipitation (24,8 g P ha -1 yr -1 ). However, the input of total P by precipitation accounted for only 0,5% of the P requirement of the trees. The deposited P may originate from distant or local sources (Attiwill & Adams, 1993; Newman, 1995; Brady & Weil, 1999). Major P sources for long-range atmospheric transport are soil-derived dust from arid and desert regions, volcanic activity, marine aerosols, combustion sources, agriculture and industry (Newman, 1995; Mahowald et al., 2008; Kopácek et al., 2011a and references therein). Local sources are pollen, plant fragments and primary biogenic aerosols (e.g. bacteria and detritus) (Newman, 1995; Mahowald et al., 2008; Kopácek et al., 2011a and references therein). According to Kopácek et al. (2011a), these local sources may significantly contribute to, or even dominate, the total atmospheric P input in forested areas away from desert regions. According to a recent review by Mahowald et al. (2008), mineral aerosols are the dominant source of total P on a global scale (82%), while primary biogenic particles (12%) and combustion sources (5%) are most important in nondusty regions. Note that on a landscape level, local deposition may also be regarded as throughfall. The amount of P in atmospheric deposition that is soluble is highly variable, ranging from 2% to 80% (Mahowald et al., 2008). Kopácek et al. (2011a) reported the average proportion of dissolved reactive P to constitute 50% of total P deposition in forested and mountaineous catchments in the Czech republic. Other studies in non-agricultural Europan sites have reported values around 40% (Persson & Broberg, 1985; Kopácek et al., 2011b). The deposition of total P in precipitation as well as in throughfall has been reported to show some seasonality, with rates being higher in spring (Kopácek et al., 2011b). This increase is generally associated with higher occurrence of pollen in spring (Kopácek et al., 2011b) Throughfall As compared to atmospheric deposition, throughfall may either be enriched or depleted in nutrients. The major reasons for increasing element concentrations in throughfall are water evaporation from the canopy, washing off of dry deposited particles, horizontal deposition and leaching of elements from the plant leaves. Uptake or microbial transformations in canopies may, on the other hand, decrease concentrations of P in throughfall compared to precipitation. In general, there are few studies investigating the fate of P when passing through the tree canopy and even fewer investigating the causes behind it. When throughfall and stemflow are collected in forests, the nutrient content per m2 ground is often higher than in the rain falling on the canopy (Newman, 1995). The studies on throughfall reviewed by Newman (1995) showed that at all but one site (all sites were in the temperate region), the rainfall gained P as it passed through the canopy. This gain of P ranged from 0,2 to 3,2 kg P ha -1 yr -1. Zhang & Mitchell (1995), investigating a hardwood forest stand in the eastern US, also found P in throughfall to be substantially higher than in rainfall. Dry deposition and (or) foliar leaching contributed up to 86% of the P input to the soil (157,6 g ha -1 yr -1 ). Kram (2010), on the other hand, found that on average 60 % of the inorganic P was retained in the canopies, when investigating the concentration of PO 4 3- in bulk precipitation and throughfall in five different forest ecosystems in central Poland. There was a clear difference between tree species, with pine and alder keeping significant amounts of inorganic P in their canopy (84%), while no such significant effects were found for birch, oak and locust. In contrast to Kram (2010), Bhat et al. (2011) found that concentrations of total P in throughfall tended to be higher in pine stands than in hardwood, wetland and mixed forest stands typical of the south-eastern US. However, the difference was not significant. When looking at average fluxes for the study period (January 2002 to August 2003), total P were often higher in rainfall than in most of the forest stands indicating an uptake of P in the tree crowns. The fluxes of total P in rainfall were similar between growing and dormant seasons. However, fluxes of P in forest stands were consistently higher (doubled) during the dormant season than during the growing season, indicating that P was taken up by canopies during the growing season and washed off during the dormant season. The seasonality in the P use by tree crowns was supported by results reported by Kopácek et al. (2009). They found that the forms of P changed considerably while passing through the canopy of six Norway spruce dominated sites in the Bohemian forest in the Czech republic. Dissolved reactive P decreased, while dissolved organic P and particulate P increased indicating microbial transformations of P in the canopy, where easily bioavailable P compounds are transformed into organic forms. The effect was most pronounced in spring and summer. Apart from the type of plant cover (Kram, 2010; Bhat, 2011), age of foliage, canopy closure and the plant s nutrient supply are factors that have been shown to affect whether nutrients in general (i.e. not P in particular) are leached or retained in the canopy (see Kram, 2010 and references therein). With regard to P, Hagen-Thorn et al. (2006) found that the amount of P leached in four deciduous forest trees in Lithuania 10

13 was related to the nutrient concentration in the foliage. Birch, which had the highest green leaf P concentrations also leached higher amounts of P than any other species. Ash, however, lost higher percentages of their P contents. The amounts of nutrients leached from the tree crowns also seemed to depend on leaf morphology. The highest amount of P, and other nutrients, per unit leaf area were leached in the ash stand, indicating that this species had high susceptibility to nutrient losses by canopy leaching. 4.2 Weathering There is not much information available with regard to the weathering of P and in many reviews of P cycling, weathering is only briefly mentioned. Most of the information here is thus based on the review by Newman (1995). Rocks break down by physical or chemical weathering. The physical weathering implies the breaking of the rock into smaller pieces, ultimately sand grains or smaller. This is carried out by frost, salt crystals, swelling of plant roots, chemical attack and other agents. Chemical weathering, on the other hand, involves the breakdown of the molecules that form the rock matrix, releasing water-soluble smaller molecules or ions. The contribution of physical weathering to the release of P is by exposing a larger surface area of particles to chemical weathering (Newman, 1995). Most of the P in rocks is contained in apatite (general formula Ca 5 (PO 4 ) 3 X where X can be F, Cl or OH), which occurs only as a minor component of rock minerals (Newman, 1995). The rate of release under field conditions is extremely difficult to measure. According to the studies reviewed by Newman (1995), rates of P release by chemical weathering may vary from 0,05 to 1,0 kg P ha -1 yr -1. However, based on the concentrations of P in rocks and assuming that the percentage of P released by weathering per year is the same for P as for the rock as a whole, Newman (1995) arrives at a possible range of P release of up to 5 kg P ha -1 yr -1. According to Newman (1995), the causes for the variation in rates of chemical weathering include: 1) rock type, 2) surface area of mineral particles per hectare of ground, 3) how much of this surface is in contact with water (higher weathering rates at higher moisture levels), 4) temperature (higher weathering rates at higher temperature) and 5) chemicals dissolved in the water, e.g. H + -ions and chelating agents produced by roots and microorganisms. Among rock types, weathering rates can be ranked as follows: quartz < aluminosilicates < carbonates < rock salts. There are no consistent differences in weathering rates between major rock types among the aluminosilicate rocks (Newman, 1995). Newman (1995) also emphasizes that if P is released more rapidly than the major rock constituents, one would expect the rate of P release to decline as weathering proceeds. This could lead to P release being slower from clay than from silt, in spite of the larger surface area of clay. The depth of rooting is likely to be important in determining how much of the P released by weathering that is able to circulate in the biomass, rather than being fixed into seconday minerals or lost by leaching (Newman, 1995). At many sites, fine weatherable material extends far deeper than the roots and according to the review by Newman (1995), more than half of the chemical weathering at some sites occurs below the root zone. Some of this will be fixed into secondary minerals, and some will be lost by leaching. 4.3 P fertilization The information given in this section refers mainly to management aspects of P additions. For further information about the biological effects of P addition, see sections 5 and 6. Most forests occur on soils that are less fertile than soils used for agricultural crops. Consequently, nutrients (especially N and P) limit growth of many forests and fertilization with P is a common practice in many parts of the world (see for example review by Fox et al., 2011). The form of P applied varies widely and range from diammonium phosphate (DAP) to ground rock phosphate and blends of N, P and potassium (K) (Fox et al., 2011). According to recently presented data for 50 years of fertilization trials on Maritime pine (Pinus pinaster) in south-western France, the form of the P fertilizer applied does not seem to have a significant impact on the outcome of P addition on tree productivity (Trichet et al., 2009). However, according to Brady & Weil (1999), using ammonium (NH 4 + ) together with P when applying the fertilizer usually greatly increases root uptake of P, especially in alkaline soils. This is probably due to an increase in the acidity of the soil when NH 4 + is added, since both root uptake of NH 4 + and nitrification are acidifying processes (Brady & Weil, 1999), which in alkaline environments may increase the solubility of P (see section 5.3.1). Generally, the P in fertilizer is subject to adsorption to or fixation by soil particle surfaces and only 10 to 15% of added P is likely to be taken up by plants in the year of fertilization (Brady & Weil, 1999). 11

14 5. P IN THE SOIL 5.1 Forms and content of P in soils The earth s crust contains on average 0,12% P (Tiessen et al., 2011). According to Brady & Weil (1999), the P content in soils range from 200 to kg P in the upper 15 cm of one hectare of soil, with an average of around kg P, while Jones & Oburger (2011) present values of total P in topsoil (0-15 cm) of between 50 and 3000 mg kg-1, depending on parent material, soil type, vegetation cover and soil management. For 18 forest soils in south-western France (mainly moorlands but also including some dunes), Achat et al. (2009b) presented numbers ranging from mg P kg-1 in litter, 16 to 109 mg P kg-1 in the top soil (0-15 cm) and mg P kg-1 in the lower soil (15-30 cm). For some other forest soils reviewed in Achat et al. (2009b), numbers are considerably higher: mg P kg-1 in litter, mg P kg-1 in the top soil (0-15 cm) and mg P kg-1 in the lower soil (10-20 cm). Generally, the P levels in soils are usually no more than one-tenth to one-fourth that of N (Brady & Weil, 1999). The bulk of the soil P exists in three general groups of compounds: 1) organic P, 2) calcium (Ca) bound inorganic P and 3) iron (Fe) or aluminium (Al) bound inorganic P (Brady & Weil, 1999). A number of organic P compounds are present in soil, but their identity and amounts are not well understood. However, three broad groups of P are known to occur in soil: 1) monoester phosphates, primarily inositol phosphates and sugar phosphates, 2) nucleic acids and 3) phospholipids (Brady & Weil, 1999; Paul & Clark, 1996; Fox et al., 2011). Of these, inositol phosphates are generally the most abundant (Brady & Weil, 1999; Jones & Oburger, 2011). However, the fraction of organic P that they constitute seems to be highly variable. Suggestions range from 10% (Brady & Weil, 1999) to roughly 40% (Paul & Clark, 1996) and more than 80% (Jones & Oburger, 2011). The nucleic acids and phospholipids together have generally been thought to make up only a few percent of the organic P in most soils (Paul & Clark, 1996; Brady & Weil, 1999; 12 Jones & Oburger, 2011), but recent studies have indicated that diester phosphates (including nucleic acids and phospholipids) may be of considerably greater importance in acid forest soils than in agricultural soils. While constituting only 10% of the organic P in agricultural soils, they were found to account for more than 50% of extractable organic P in acid forest soil in a study in British Columbia (Fox et al., 2011). Diester phosphates are believed to be less tightly sorbed to soil colloids than monoesters and may thus be more susceptible to hydrolysis by phosphatases present in the soil (Fox et al., 2011). Of the inorganic P, Ca compounds predominate in most alkaline soils, while Al and Fe forms are more important in acid soils (Brady & Weil, 1999). At higher ph, the Ca phosphates are quite stable and very insoluble. As soil ph decreases they become more soluble, hence they tend to dissolve and disappear from acid soils. For Al and Fe phosphates, the opposite is true (Brady & Weil, 1999). Of the common Ca compounds containing P, the apatite minerals are the least soluble. Some apatite minerals are so insoluble that they even persist in highly weathered (i.e. acid) soils. The simpler mono- and dicalcium phosphates, on the other hand, are more readily available for plant uptake (Brady & Weil, 1999). Fe- and Al-hydroxy phosphates have very low solubilities in acid soils. However, they become more soluble as soil ph rises. The relative amounts of organic and inorganic P vary greatly from soil to soil. Brady & Weil (1999) presented numbers where the organic fraction constituted 20 to 80% of the total P in the surface soil horizons, while Paul & Clark (1996) stated that it may range from 5 to 95% and Fox et al. (2011) presented a range between 20 and 90%. Lambers (1998) suggested that between 30 and 70% of all P present in agricultural soils is in the organic form, while in nutrient poor grassland and forest soils it may constitute as much as 80 to 95% and in organic tundra soils 99%. Jones & Oburger (2011) suggested that between 30 and 65% of total P is present as organic P in mineral soils, while in organic soils (>20-30% or-

15 ganic matter), this fraction may approach up to 90% of the total P. Achat et al. (2009a), using isotopic dilution methods to assess the availability of P, found that dead soil organic matter contained 77% of total P in a sandy low P-sorbing pine forest soil in south-western France, while microbes contained 17% of total P. Similar values for the microbial content of P was reported by Khanna et al. (2007). In their review study, 13% and 23% of total P in the mineral soil and the forest floor were reported to be found in microbial biomass. Achat et al. (2009a) concluded that inorganic P sorbed to the solid phase represented a small but rapidly available pool of P, while P in dead soil organic matter represented a considerably larger but slowly available pool. In another experiment, combining isotopic dilution and extraction methods, Achat et al. (2009b) evaluated the contribution of organic P at various depths in the soil at 18 forest sites in the south-western part of France. The contribution of organic P to P availability (in relation to the contribution of inorganic P) was predominant in litter (average 79%, range 59-88%), less important in top soil horizons (average 56%, range 36-88%) and rather small at depths below 30 cm (average 45%, range 21-60%). This was due to a decreasing proportion of organic P and a higher stability of organic compounds with soil depth, as well as an increasing proportion of diffusive P (ionic P species that can be transferred from the solid phase to the soil solution due to a gradient of concentration) with depth. According to Achat et al. (2009b), the relative contribution of organic P did not necessarily increase with increasing amounts of total stocks of organic P. Owing to a very low amount of diffusive P in the top soils in dry sites, the relative contribution of organic P was higher in these sites than in humid and mesic sites, despite a lower overall organic P fraction. The major P pool in hardwood stands has been found to be in the mineral soil (Zhang & Mitchell, 1995). In contrast with the general view that boreal forests have a geochemical P sink in the mineral soil, Giesler et al. (2002) found that such a sink was also present in the humus layer in ground water discharge areas. This was according to the authors due to a redistribution of Al and Fe, probably as a consequence of an upward water flow. Beck & Elsenbeer (1999) showed that the organic fraction at various depths may vary with stand age. When investigating two beech forest Spodosols in the southern Alps, they found that the distribution of organic P did not vary much with depth in the younger stand (46 years), while there was a clear gradient with a larger fraction of organic P in the upper layers of the soil in the older stand (108 years). Fertilization with P generally increases the content of P in the soil, although the availability and in which part of the soil the P ends up may vary substantially (see for example Finzi, 2009; Prietzel & Stetter, 2010; Groffman & Fisk, 2011). For agricultural soils in Sweden, Börling et al. (2004) found that 99% of the P added was sorbed by soils when using P sorption index 1 (PSI1, i.e. addition of 19,4 mmol P kg -1 soil), despite the soil having received kg P ha -1 yr -1 for more than 30 years. 5.2 Mineralization and immobilization of P P held in the organic form can be mineralized and immobilized by the same general processes that release N from soil organic matter. Forest floors have been found to be sinks, as well as sources, of P (Fox et al., 2011). Should organic residues low in P but high in C and other nutrients be added to a soil, microbes will immobilize the P in their biomass. According to Brady & Weil (1999), net immobilization of soluble P occur if residues added to the soil have a C/P ratio greater than about 300:1, while net mineralization occur if the ratio is below 200:1. However, Moore et al. (2006; 2011) presented considerably different ratios for Canadian temperate to subarctic forests. Moore et al. (2006) found a strong relationship between the initial P concentration in litter and the pattern of P release, as litters with high initial P concentrations lost P and those with low concentrations retained P, with a critical C/P quotient for release of Moore et al. (2011) suggested that net P loss likely occurred at C/P quotients between 800 and Osono & Takeda (2004) investigated the influence of lignin to P ratios (L/P) as indicators of P dynamics in litter in a cool temperate forest in Japan. The L/P ratio at which mobilization began was The critical values of the L/P ratios showed convergent trends among litter types as compared to their initial values, and approached those of the underlying humus layer (Osono & Takeda, 2004). That the average and the range of C/P fell and narrowed during decomposition was also found by Moore et al. (2011). However, the ratios remained substantially larger than in the decomposer organisms. According to Brady & Weil (1999), mineralization of organic P in soils in temperate regions typically release between 5 and 20 kg P ha -1 yr -1. These values can be compared to the annual uptake of P by crops, trees and grasses, which generally ranges from 5 to 30 kg P ha -1 yr -1 (Brady & Weil, 1999). In spite of the fact that P may limit productivity of forest ecosystems, much less attention has been given to P mineralization as compared to N mineralization (Fox et al., 2011). According to Brady & Weil (1999), mineralization of organic P is subject to many of the same influences that control the general decomposition of soil organic matter, such as temperature, moisture and availability of N. Because of the limited information available with regard to the mineralization of P, a description of the main factors controlling decomposition in general follows (sections to 5.2.5). Where information was available, factors of specific importance for the mineralization of P have been added. A number of studies have also demonstrated significantly increased rates of organic P mineralization in response to soil phosphatase activity. For further information about this, see George et al. (2011) and references therein Effects of climate Climate is the most important factor determining decomposition of plant litter (Swift et al., 1979; Aerts, 1997; Berg & McClaugherty, 2003), and variation in decomposition rates between climatic regions has been shown to reflect variations in macroclimate, primarily soil moisture and temperature (Meentemeyer, 1978). The effects of soil temperature and soil moisture are generally both positive, at least within a certain range (Donnelly et al., 1990; Mahli et al., 1999 and references therein) and for easily decomposable pools of SOM (Melillo et al., 2002; Pendall et al., 2004 and references therein). Labile substrates are generally regarded as making up approximately 10% of the total SOM pool. Whether the later stages of decomposition and SOM are sensitive to temperature is unclear, with some studies suggesting increasing sensitivity with decreasing quality (Fierer et al., 2005; Knorr et al., 2005) while others suggest no effect (Giardina & Ryan, 2000; Fang et al., 2005). Recently, Karhu et al. (2010) showed that the temperature sensitivity of decomposition increases substantially, from the youngest annually cycling fraction (Q 10 <2) to a decadally cycling fraction (Q 10 =4,2-6,9), but decreases again to a centennially cycling fraction (Q 10 =2,4-2,8) in two upland boreal forest soils. Applying the same temperature sensitivity of decomposition for all SOC fractions may thus, according to Karhu et al. (2010), give a biased picture of future SOC cycling. Davidson & Janssens (2006) raised the issue of intrinsic and apparent temperature sensitivity. It is well known that SOM consists of more or less of a soup of thousands of different organic C compounds, each with its own inherent kinetic properties. The inherent kinetic properties based on molecular structure and ambient temperature is called intrinsic temperature sensitivity of decomposition by Davidson & Janssens (2006). However, in addition to the complex structures of the organic matter, the enzymes for decomposition may be physically or chemically excluded from many of the organic C substrates within the heterogenous soil environment, causing substrate limitation at the reaction microsite. The observed response to temperature under these environmental constraints is called apparent temperature sensitivity by Davidson & Janssens (2006). The apparent temperature sensitivity may thus be much lower than the intrinsic temperature sensitivity of the substrate. On the other hand, it may also be higher (if a temperature sensitive process alleviates an environmental constraint to decomposition, then the subsequent increase in substrate availability could result in the apparent temperature sensitivity temporarily exceeding the intrinsic temperature sensitivity of the substrate). Davidson & Janssens (2006) lists a number of factors that can temporarily or indefinitely affect apparent temperature sensitivities of decomposition: 1) Physical properties: Organic matter may become physically protected in the interior of soil aggregates. They can also be physically protected from degradation by water-soluble enzymes if they have low water solubility, or if they occur in hydrophobic domains of humified organic matter. 2) Chemical protection: Organic matter may become adsorbed onto mineral surfaces, through covalent or electrostatical bonds, thus chemically protecting it from decomposition. 3) Drought: Drought reduces the thickness of soil water films, thus inhibiting diffusion of extracellular enzymes and soluble organic C substrates. 13

16 4) Flooding: Flooding slows oxygen diffusion to decomposition reaction sites. 5) Freezing: The diffusion of substrates and extracellular enzymes within the soil is extremely slow when soil water is frozen. Each of these environmental constraints affects decomposition reaction rates, directly or indirectly, by decreasing substrate concentrations at enzymatic reaction sites. In some cases, the influence of climate on decomposition also seems to depend on vegetation type. For needle litter of Scots pine, it has been demonstrated that the dominant rate-regulating factor influencing the decomposition rate, both the annual variation at one site (Jansson & Berg, 1985) and within a 2000 km-long climatic transect (Berg et al., 1993) was the climate. In a corresponding transect of Norway spruce forests, on the other hand, no effect of climate was detected (Berg et al., 2000). In a study by Vanhala et al. (2011), transplantation of organic surface horizons of boreal soils into warmer regions altered both the microbial community composition and vegetation (and thus the C source that the microbes feed on), without altering the temperature sensitivity of decomposition. With regard to P, Sardans & Penuelas (2007), investigating the effects of a 6-year-long drought in a Mediterranean Quercus ilex forest, found that drought slowed down the release of P from organic to inorganic forms, leading to an accumulation of soluble organic forms and, consequently, to an increase in total soluble P. Furthermore, disturbances such as harvesting may also significantly affect the P dynamics of the forest floor, due to its impact on temperature and moisture (Fox et al., 2011). Freeze-thaw cycles have also been found to influence soil nutrient mineralization, including the mineralization of P (Freppaz et al., 2007). Both concentrations of total dissolved P (TDP) and dissolved organic P (DOP) were increased after the freeze-thaw event in four and three out of four soils, respectively (Freppaz et al., 2007). The soils were of differing types and collected in the western Alps. The relative contribution of DOP to TDP after the freeze-thaw event remained significant and greater than 50% in all soils. There were few effects of an increased number of freeze-thaw events (Freppaz et al., 2007). That soil freezing may also have significant effects on soil solution concentrations of inorganic P was shown by Fitzhugh et al. (2001). The quantities that were mobilised from the forest floor (15-32 mol ha -1 yr -1 ) in a northern hardwood forest in the US were significant relative to the available P pool (22 mol ha -1 ) and the rate of net P mineralization in the forest floor (180 mol ha -1 yr -1 ). However, the elevated leaching of inorganic P from the forest floor was coupled with enhanced retention in the mineral soil and the inorganic P was thus largely removed from the solution. According to the authors, increased fine root mortality was most likely an important reason for the increase in leaching of mobile inorganic P from the forest floor, but other factors such as decreased P uptake by roots and increased physical disruption of soil aggregates were also suggested to contribute. With regard to P mineralization in cold arctic and alpine soils, it has been shown that net P mineralization may be low, or even negative, during the growing season (Weintraub, 2011). Recent studies have demonstrated, however, that the activity of both microbial biomass and enzymes may be relatively high in snow-covered soils, with significant implications for P cycling. Several studies reviewed in Weintraub (2011) demonstrate a crash in microbial biomass and a subsequent rapid release of P after the soil thaws in spring. Chapin et al. (1978) estimated that the spring pulse of dissolved inorganic P from such a crash represented as much as 30% of annual P uptake in a wet sedge meadow on the arctic coastal plain of Alaska. According to Weintraub et al. (2011), the P in this pulse may be taken up immediately, or may enter the exchangeable P pool. However, there are also studies showing no indication of such a pulse release of P (Weintraub, 2011 and references therein). Weintraub (2011) thus caution that while this pattern has been observed in both arctic and alpine soils, it does not appear to be a universal feature of cold soils. Why it occurs in some soils and not in others is according to Weintraub (2011) still unclear. Variations in cold-adaptability of the soil microbial community, soil temperature, moisture content at the time of freezing, snowfall, depth of snowpack and rate of soil thaw have all been suggested as possible factors influencing the presence (or absence) of the pulse (Weintraub, 2011). 14

17 5.2.2 Effects of N Apart from climate, the quality of litter in terms of its susceptibility to attack by decomposers seems to be the most important factor affecting the rate of decomposition (Attiwill & Adams, 1993). Traditionally, the C/N ratio has been regarded as a good indicator of decomposability. However, nowadays, the initial concentration of N and the lignin/n ratio are often considered as the best predictors of litter decomposition rates (Kimmins, 1997; Berg, 2000). It has been suggested that decomposition can be divided into two phases; an early stage in which climate as well as concentrations of the major nutrients and water soluble substances has a clear influence on decomposition rate, and a later phase where the decomposition of lignin dominate over the influence of nutrients and thus rule the decomposition of litter (Berg & Staaf, 1980; Taylor et al., 1989; 1991). The mechanism was further explained in Berg (2000): In fresh litter, the degradation process is dominated by easily soluble C compounds. Consequently, the amount of C is high in relation to macronutrients such as N. Net N immobilization may thus occur, resulting in a deficit of N in relation to C. Consequently, initial litter decomposition rates respond positively to increased N availability. With the disappearance of celluloses, the concentration of the more recalcitrant compound, lignin, increases, and in partly decomposed litter the degradation rate of lignin determines the decomposition rate of the whole piece of litter (which is now turning into SOM). The suppressing effect of lignin on litter massloss rates can be described as a linear relationship (Berg & Lundmark, 1987), which, for pine litter may start already at approximately 20-30% mass loss. In contrast to the early phase, high N concentrations will now have a rate-retarding effect on lignin degradation and thus on the litter decomposition. This retardation has been shown to be due to suppression of lignolytic enzymes in white rot fungi, but can also be a result of the formation of chemically stable recalcitrant compounds, which are formed when low-molecular N that reacts with lignin remains (Berg, 2000). Another explanation was suggested by Chapin et al. (2009). In the presence of adequate N, the priming effect is diminished as microorganisms preferentially utilize C-rich substrate additions. Accordingly, the end result is that increases in soil N may promote SOM conservation (Berg & Meentemeyer, 2002; Craine et al., 2007). When N availability is low, on the other hand, microorganisms mine SOM for N, increasing the SOM- C respired. Chapin et al. (2009) also suggested that the differences in C sequestration in response to N addition may depend on whether the largest effect of N in soils is to increase the non-biological formation of recalcitrant SOM (reduces decomposability) or to increase growth and metabolism of soil decomposers (N stimulation of decomposition). Berg (2000) presented data showing that different tree species vary considerably with regard to their content of lignin and N. Scots pine needle litter, for example, has simultaneously low concentrations of both N and lignin, whereas lodgepole pine litter has low N but high lignin concentrations. Norway spruce has higher N concentrations than those of the two pine species, but lignin concentrations that are in between, while birch leaves has lignin concentrations similar to those of the spruce needles and generally higher N concentrations. The leaves of common beech are extreme, with very high concentrations of both N and lignin Effects of nutrients other than N In northern coniferous forests, the rate of dry-matter loss in decomposing litter of Scots pine was found to be more dependent on the initial P concentration of the litter than on the initial N concentration (Meentemeyer & Berg, 1986), and P was immobilized to a much greater extent than was N (Staaf & Berg, 1982). In coherence with these studies, Polglase et al. (1992a,b) showed that release of P from decomposing needles in pine plantations was strongly and positively related to the needle concentration of inorganic P (which increased with fertilization). Furthermore, Berg (2000) showed that high initial concentrations of manganese (Mn) and Ca in litter had an opposite effect to N, i.e. high concentrations of these two nutrients were correlated to further decomposition of the litter. The response of Mn was linear in a concentration range of Mn from 0,05 to 4,71 mg g -1. Ca has been found to support the growth of a white-rot fungal species of importance for decomposition (Berg, 2000), while Mn is essential for the activity of Mn-peroxidase, a lignin-degrading enzyme (Perez & Jeffries, 1992). Mn is also involved in the regulation of other lignolytic enzymes (Archibald & Roy, 1992; Perez & Jeffries, 1992) Effects of aluminium and ph Increasing soil ph generally has a positive effect on decomposition of both labile and more recalcitrant litter compounds (Benner et al., 1985; Donnelly et al., 1990; Kreutzer, 1995; Kurka et al., 2001). An explanation for this effect was provided by Donnelly et al. (1990), who suggested that the extracellular enzymes involved in the degradation of resistant compounds were inhibited by acidic conditions. Other studies have shown that the acidity of the soil may also influence the biomass and the composition of the decomposer community (Kreutzer, 1995), which affects not only the decomposition rate but also the formation of resistant compounds (Nierop & Verstraten, 2003). For example, fungi, mites and springtails are less adversely affected by low ph and generally dominate under acid conditions. Since decomposition tends to be slow when it is primarily fungal, it is thus generally slower under acid conditions. In less acid soils, earthworms and bacteria dominate, speeding up the decomposition rate (Kimmins, 1997). With regard to P, acid-mist treatment (a mixture of H 2 SO 4 and NH 4 NO 3 with ph 2,5) of a Sitka spruce (Picea sitchensis (Bong.) Carr) stand in Scotland significantly lowered the availability of P (water-soluble P content in forest litter, labile inorganic P content and phosphate concentrations in soil solution) in soil layers beneath acid-mist treated trees compared with control trees (Carreira et al., 1997). The labile organic P content, on the other hand, increased. According to Carreira et al. (1997), the effects were probably a consequence of the decrease in ph and base saturation in soils under trees exposed to the acid mist treatment, resulting in an increased P sorption capacity of the surface soil and a slowing of the P cycling (as indicated by the accumulation of organic P). Closely and mutually related to soil ph is aluminium (Al). Increasing Al concentrations were found to cause a decrease in the decomposition rate of cellulose. This was partly due to reduced enzymatic activity, as in the case of acidity, but also to reduced substrate availability, due to binding of polysaccharides with Al-hydroxide which resulted in the formation of more stable and less water soluble compounds (Miltner & Zech, 1998). Al has also been found to have a negative influence on the degradation of organic acids, as a consequence of the formation of complexes with citric acid (Jones et al., 2001). However, the negative effect was evident only at concentrations much higher than those normally found in forest soils, and Jones et al. (2001) emphasized that roots generally produce extensive amounts of organic acids when Al is present in the soil solution, and that these acids are usually rapidly mineralized by the soil microorganisms Effects of parent material Parent material also influences P mineralization rates. Lopez-Hernandez & Nino (1993) showed that for soils with the same taxonomic classification (in their study Typic Haploborolls), mineralization rates were significantly higher in fine-textured soils compared with more coarse textured ones, corresponding to a higher content of organic P and microbial P in the former. 5.3 Adsorption and desorption of P The fixation or adsorption of phosphate ions on the surface of soil particles exerts a strong control on the concentration of inorganic P in soils (Brady & Weil, 1999). Freese et al. (1995) investigated the phosphate sorption kinetics in acid soils and found that by combining an equilibrium model and a kinetic model, the total sorption of P could be described by the following equation: F (c,t ) = F m (ki) 1+ (ki) where F(c,t)=total sorption Fm=sorption maximum I=exposure k=rate coefficient K=adsorption constant c=concentration in bulk solution Kc 1+ Kc 15

18 The first term on the right-hand side (Fm) represents the sorption maximum. In order to reach the sorption maximum, relatively high concentrations and/or long reaction times are required. The second term on the right hand side is the kinetic part and its value varies between 0 and 1 depending on the extent of the exposure. The combination of the first two terms on the right-hand side of the equation can be interpreted as the availability of the binding sites which depends on the exposure of the system with phosphate. The third term expresses that the fraction of the available binding sites, occupied with phosphate, depends on the phosphate concentration in the bulk solution. Freese et al. (1995) found that the Fm values were strongly correlated with the sum of oxalate extractable Fe plus Al. All soils investigated in Freese et al. (1995) (sandy as well as non-sandy) could be reasonably well described by the model with a S-shaped curve clearly visible when plotting fractional saturation of phosphate-sorption capacity (F(ct)/Fm) as a function of the exposure (ln I(c,t)). The particular types of reactions that fix P into relatively unavailable forms differs from soil to soil, but are generally closely related to ph (Brady & Weil, 1999). Other factors that tend to control the sorption reactions in soils are soil age and contents of organic matter and clays Effects of ph As referred to above, ph has a major influence on the solubility of P containing compounds (Lindsay, 1979; Brady & Weil, 1999). In acid soils, these reactions generally involve Al and Fe (and to a smaller extent Mn) either as dissolved ions or as hydrous oxides. Many soils contain such hydrous oxides as coatings on soil particles and as interlayer precipitates in silicate clays (Lindsay, 1979; Brady & Weil, 1999). Al H 2 PO H 2 O 2H + + Al(OH) 2 H 2 PO 4 While the phosphate ion is soluble, the reaction product is insoluble. In alkaline and calcareous soils, the reactions primarily involve precipitation as various Ca-phosphate minerals: Ca(H 2 PO 4 ) 2. H2 O + 2H 2 O 2(CaHPO 4. 2H 2 O) + CO 2 (g) CaCO3 CaCO3 2(CaHPO 4. 2H 2 O) + CO 2 (g) Ca 3 (PO 4 ) 2 + CO 2 (g) + 5 H 2 O Monocalcium phosphate (Ca(H 2 PO 4 ) 2 ) is soluble and dicalcium phosphate (Ca(H 2 PO 4 ) 2 ) slightly soluble, whereas tricalcium phosphate (Ca 3 (PO 4 ) 2 ) has very low solubility. Apart from reacting with Ca, P may also be adsorbed to the Fe impurities on the surfaces of carbonates and clays in alkaline soils (Lindsay, 1979; Brady & Weil, 1999). At moderate ph, adsorption on the edges of kaolinite or on the Fe-oxide coating on kaolinite clays plays an important role (Lindsay, 1979; Brady & Weil, 1999). The greatest degree of P fixation occurs at very low and very high soil ph. As ph increases from below 5.0 to about 6.0, the Al and Fe phosphates become somewhat more soluble. Also, as ph drops from greater than 8.0 to below 6.0, the Ca phosphates increase in solubility (Brady & Weil, 1999). Consequently, as a general rule, plant availability of P is at its highest when soil ph is between 6.0 and Effects of clay content Most of the compounds with which P reacts are in the finer soil fractions (Brady & Weil, 1999). Therefore, if soils with similar ph values and mineralogy are compared, P fixation tends to be more pronounced in those soils with higher clay contents (Brady & Weil, 1999). Furthermore, some clay minerals are more efficient in fixing P than others. In general, clays with high anion exchange capacity have a greater affinity for phosphate ions. For example, oxides of Fe and Al, such as gibbsite and goethite, strongly attract and hold P ions. Also, kaolinite has a great P fixing capacity, while the 2:1 clays of less-weathered soils generally have relatively low capacity to bind P (Brady & Weil, 1999). The large quantities of Fe- and Al-oxides and 1:1 clays present in many soils make possible the fixation of extremely large amounts of P by these reactions. Brady & Weil (1999) suggest the following order for the P-fixing capacity of soil components (increasing extent and degree of fixation): 2:1 clays << 1:1 clays < carbonate crystals < crystalline Al, Fe and Mn oxides < amorphous Al, Fe and Mn oxides, allophane According to Brady & Weil (1999), the soil components are to some degree distributed among soils in relation to soil taxonomy. Vertisols and Mollisols are generally dominated by 2:1 clays and have low P fixing capacity, while Ultisols and Oxisols are dominated by Fe- and Al-oxides with higher P-fixing capacity. Andisols, characterized by large quantities of amorphous oxides and allophane, have the greatest P-fixing capacity, and their productivity is often limited by this property (Brady & Weil, 1999). However, Beck & Elsenbeer (1999) caution against a reliance on soil taxonomy as a framework for interpretation of soil P forms. They found, when comparing three Spodosols from the southern Alps, that they differed considerably with regard to the crystallinity of free Al, the presence of amorphous compounds and their respective distributions with depth. Although the exact mechanisms are not entirely clear, H 2 PO 4 - ions are known to react with Fe- and Al-surfaces in several different ways, resulting in different degrees of P fixation. Brady & Weil (1999) describes the ways by which inorganic P may react in acid soils in the following way: 1. Precipitation The dissolved ions of Al or Fe and phosphate react and forms precipitates. Freshly precipitated Al, Fe and Mn phosphates are relatively available, although over time, they may become increasingly unavailable. 2. Anion exchange reactions The phosphate ion may become attracted to positive charges that develop under acid conditions on the surfaces of Fe- and Al-oxides and the edges of kaolinite clays. In this state it may be replaced by other anions, such as OH - and SO 4 2- or organic acids (R-COO - ), in the reversible process of anion exchange. Consequently, it is available (although slowly) to plants. Availability of such adsorbed H 2 PO 4 - may be increased by liming of soil to increase hydroxyl ions or by adding organic matter to increase organic acids capable of replacing H 2 PO Reactions with Al- or Fe-oxide surfaces Alternatively, H 2 PO 4 - may replace a structural hydroxyl and become chemically bound to the oxide or clay surface. This reaction, although reversible, generally binds the phosphate too tightly to allow it to be replaced by other anions. Plant availability is thus low. 4. Formation of a stable binuclear bridge Over time, a second oxygen of the phosphate ion may replace a second hydroxyl so that the phosphate becomes chemically bound to two adjacent Al och Fe atoms on the hydrous oxide surface. It is now an integral part of the oxide mineral and the likelihood of its release back to the soil solution is very small. As more time passes, the precipitation of additional Fe- or Al-hydrous oxides may bury the phosphate deep inside the oxide particle. Such phosphate is called occluded and is the least available form of P in most acid soils. In more alkaline soils, soluble H 2 PO 4 - reacts quickly with Ca to form a sequence of products of decreasing solubility. Brady & Weil (1999) describe the reactions taking place when highly soluble monocalciumphosphate [Ca(H 2 PO 4 ) 2 H 2 O] is added to soil in the following way: The highly soluble monocalciumphosphate rapidly reacts with Ca in the soil to form first dicalcium phosphate [CaHPO 4 2H 2 O] and then tricalcium phosphate [Ca 3 (PO 4 ) 2 ]. The solubility of these compounds, and consequently the plant availability of P, decreases as the P changes from the H 2 PO 4 - ion to tricalcium phosphate [Ca 3 (PO 4 ) 2 ]. Although already relatively insoluble, the latter may undergo further reactions to form even more insoluble compounds, such as various forms of apatites. These are thousands of times less soluble than freshly formed tricalcium phosphates Effects of organic matter Organic matter generally has low capacity to fix phosphate ions (Brady & Weil, 1999). Instead, soils rich in organic matter, especially active fractions of organic matter, exhibit low levels of P fixation for other reasons. The presence of large humic molecules can mask the P-fixation sites on clay and metal hydrous oxide particles by adhering to them, thus preventing them from reacting with the P ions. Organic acids produced by plant roots and microbial decay processes can serve as organic anions, and thus compete with P ions for fixation sites on surfaces of clays and hydrous oxides. Certain organic acids, and similar organic compunds, may also entrap reactive Al and Fe in stable organic complexes, so called chelates. Once chelated, the metals are unavailable for reaction with P ions in solution (Brady & Weil, 1999). 16

19 5.3.4 Effects of age In both acid and alkaline soils, P tends to undergo sequential reactions that produce P containing compounds of lower and lower solubility (Brady & Weil, 1999). Therefore, the longer that P remains in the soil the less soluble and therefore less plant-available it tends to become. According to Brady & Weil (1999), the effect of aging is due to factors such as the regularity and size of crystals in precipitated phosphates, more permanent bonding of adsorbed phosphates into the Ca-carbonate or metal oxide particles, and the extent to which the sorbed phosphate is buried as surface precipitation reactions continue. Freshly precipitated hydroxy phosphates may thus be slightly soluble because they have a great deal of surface area exposed to the soil solution. The P contained in them is at least initially somewhat available to plants. Over time, as the precipitated hydroxy phosphates age, they become less and less soluble and the P in them becomes almost completely unavailable to most plants (Brady & Weil, 1999). 17

20 6. P IN THE PLANTS 6.1 P acquisition and uptake Availability of P in the soil solution Both organic P and inorganic P contribute P to the soil solution, but most of the P in each group is of very low solubility and not readily available for plant uptake (Brady & Weil, 1999). Even when soluble sources of P are added to soils, in the form of fertilizers or manure, they are usually rapidly fixed and, in time, form insoluble compounds (Brady & Weil, 1999). The latter is the reason for the substantial build-up of plant unavailable P-reserves in agricultural soils. The concentrations of P in soil solution usually range from 0,001 mg l -1 in very infertile soils to about 1 mg l -1 in rich, heavily fertilized soils (Brady & Weil, 1999). In most soils, the amount of P available to plants from the soil solution at any one time seldom exceed 0,01% of the total P in the soil (Brady & Weil, 1999). In natural ecosystems, mineralisation of soil organic P is thought to provide the major proportion of P to plants (George et al., 2011). Jonard et al. (2009) demonstrated, by using labelling of isotopically exchangeable phosphate ions, that the forest floor provided 99,1% of the P supply to seedlings of maritime pine (Pinus pinaster) in a pot experiment. After 130 days, the uptake from mineral soil was still insignificant, while the P content of seedlings grown with access to the organic layer had increased tenfold with respect to the initial P content. The higher uptake from the forest floor rather than the mineral soil is, according to the authors, probably due to the lower ability of the forest floor to retain inorganic P and subsequently, more inorganic P is maintained in solution. In a field study using radioisotope techniques (32P and 33P), Brandtberg et al. (2004) reported that the contribution of the forest floor to the supply of P was 93 and 95% for spruce and birch. In acid mineral soils with high retention capacity of P, the accumulation of decomposing litter may constitute an alternative medium that allow close coupling between of litter decomposition and root uptake (Jonard et al., 2009). Brady & Weil (1999) suggested that the easily soluble fraction of soil organic P is often the most important factor in supplying P to plants in highly weathered soils (eg. Ultisols and Oxisols), although the total organic matter content of these soils may not be especially high. However, the inorganic P in these soils is generally too insoluble to contribute much to plant nutrition (Brady & Weil, 1999). In contrast, it appears that the more soluble inorganic forms play the biggest role in less weathered soils, even though these generally contain relatively high amounts of soil organic matter (Brady & Weil, 1999). There is some evidence that where organic P accumulates in areas with relatively high moisture availability, availability of P is also higher (Weintraub, 2011). In rocky alpine habitats, on the other hand, plant litter may be transported off site, increasing the dependance on bed-rock derived nutrients (Arnesen et al., 2007) Forms of P taken up by plants Plant roots absorb P dissolved in the soil solution, mainly as orthophosphate (preferentially H 2 PO 4 - ) (George et al., 2011; Jansa et al., 2011). Although the organic P is important for plant uptake of P, there is according to George et al. (2011) no evidence for direct uptake of dissolved organic P compounds by plants. Inorganic P (preferentially H 2 PO 4 - ) is also the only form taken up by mycorrhizal fungi (Jansa et al., 2011). Although various organic forms of P have been reported as potentially utilizable by some microorganisms, Jansa et al. (2011) concluded that their enzymatic cleavage to inorganic P occurs before the cross-membrane uptake, in the close vicinity of the microbial cells Root uptake Uptake by plant roots require not only that the phosphate ions be dissolved in the soil solution, but also that they move from the bulk soil to the surface of the root. This movement, usually over a distance of a few millimetres to a centimeter, takes place primarily by physical diffusion (Brady & Weil, 1999). However, because phosphate ions are very strongly adsorbed by soil particles diffusion to the root may be very slow. Since plants are adapted to a low availability of P in the soil, they have developed a number of different strategies to increase their uptake efficiency of P (reviewed in Trolove et al., 2003 and Vance et al., 2003). These mechanisms can roughly be divided into two groups: (i) mechanisms that increase the plants uptake capacity of P and (ii) mechanisms that change the equilibrium between different P compounds in the soil, thereby increasing the soil availability of P. According to Fox et al. (2011), many of these mechanisms seem to operate on several levels and influence the availability and uptake of both organic and inorganic P. The responses are coordinated by a small number of regulatory systems controlled by both the P status in the shoot and the availability of inorganic P in the rhizosphere (see review by George et al., 2011) Mechanisms that increase the plants uptake capacity The mechanisms that increase the plant s uptake capacity include changes in the root architecture, such as extensive lateral root formation and formation of root hairs (Drew, 1975; Lynch, 1995; Gahoonia et al., 1997; Lynch & Brown, 2001), as well as physiological changes of the roots, for example upregulation of P membrane transport systems in response to P deficiency (Caldwell et al., 1992; Schachtman et al., 1998). Root traits, of course, promote uptake of all ions, but are critical for P because of the low diffusion coefficient for P in soil. Modelling studies have shown that for immobile nutrients such as P, kinetic uptake parameters are of considerably less importance than are root traits (Lambers et al., 1998). Although root traits are important, the most important mechanisms for increasing the plant uptake capacity may be the formation of microbial associations, in particular associations with mycorrhizal fungi. In soils low in available P (and N), many species of plants could barely survive without mycorrhizal assistence in obtaining P (and N) (Brady & Weil, 1999). Like root hairs, mycorrhizas increase the roots absorptive surface (Paul & Clark, 1996; Jansa et al., 2011) and according to Lambers et al. (1998), the effective root length of the mycorrhizal association may increase one hundred-fold or more per unit root length. Mycorrhizal hyphae may also access smaller soil pores than the plant roots (Paul & Clark, 1996; Lambers et al., 1998; Jansa et al., 2011), can possibly compete more efficiently than plants with other microbes (Lambers et al., 1998) and can take up P at lower concentrations in the soil solution than can plant roots (Paul & Clark, 1996). Like plant roots (see section ), mycorrhizal fungi may also exude phosphatases and organic acids to enhance the availability of P in the soil (Paul & Clark, 1996; Lambers et al., 1998; Cairney, 2011; Fox et al., 2011; Jansa et al., 2011; Plassard et al., 2011). Many ectomycorrhizal fungi are also thought to be involved in the mineralization of organic nutrients and bioweathering of recalcitrant inorganic nutrients from carbonates, micas and apatites (Jansa et al., 2011). Several studies showing heavy colonization of ectomycorrhizal fungi of mineral grains (Hagerberg et al., 2003), evidence of fungal-induced weathering of apatite (Smits et al., 2008) and the ability of ectomycorrhizal fungi to produce a range of organic polymer-degrading enzymes that can attack molecules such as chitin, lignins, polyphenols and tannins as well as P-containing molecules (see for example review by Read & Perez-Moreno, 2003) have been presented. However, according to the reviews by Cairney (2011) and Jansa et al. (2011), other soil microorganisms are likely to have been involved in the release of nutrients from minerals and organic compounds, and the contribution of mycorrhizal fungi to these processes remains poorly quantified. What is clear, however, is that ectomycorrhizal fungi significantly modify soil conditions in the rooting zone (aggregation, wettability, biological activity), modulate intra- and interspecific competition within the plant community, affect soil microbial communities and have the saprotrophic capacity to intervene in microbial mobilization-immobilization cycles and to sequester P (and N) from the organic complexes formed during the decomposition process (Jansa et al., 2011). In forest ecosystems, the latter is, according to Jansa et al. (2011), the dominant source of P. According to Smith & Read (2008), it is the P concentrations in the plant rather than in the soil that determine the extent of mycorrhizal colonization of plants roots and Cairney (2011) concluded that host P demand seems likely to be a key driver of processes governing movement of P in ectomycorrhizal mycelia in the soil and P transfer to the plant. Jansa et al. (2011) emphasizes, however, that the extent of root colo- 18

21 nization by different fungi and the magnitude of benefits depend on a broader environmental context and not only on the P availability. These additional factors include soil N levels, plant age, growth rate, the identity of both plant and fungal species and, with regard to regeneration, also seed size and nutrient reserves contained in the seed (Jansa et al., 2011) Mechanisms that change the equilibrium between different P compounds in the soil Among the plant mechanisms that change the equilibrium between different P compounds in the soil, exudation of compounds that solubilize P is one of the most important. Among these compounds are protons, which may decrease the rhizosphere ph and thus alter the solubility of many different P compounds (Marschner, 2003), phosphatases (Tarafdar & Claasen, 1988; Duff et al., 1994; Asmar et al., 1995), and low molecular organic acids (Attiwill & Adams, 1993; Jones & Darrah, 1994; Jones, 1998; Ryan et al., 2001). Phosphatases are a group of enzymes that hydrolyse soil organic P, releasing inorganic P that may be absorbed by the roots. The production of phosphatases is generally related to P demand (Weintraub, 2011 and references therein), enhanced by a low P supply to the plants (Lambers et al., 1998) and may occur at soil temperatures as low as -20 C (Weintraub, 2011). Depletion of various pools of extractable organic P from the rhizosphere has been linked with greater phosphatase activities around plant roots (George et al., 2011; Nannipieri et al., 2011). However, according to George et al. (2011), the relative contribution of extracellular phosphatases derived from roots and from microorganisms in the utilisation of soil organic P is unclear because the number and activity of bacteria and fungi are greater within the rhizopshere than in the bulk soil. Furthermore, Fransson & Jones (2007) showed that at concentrations of organic P typical of fertile grassland soils, phosphatase activity did not limit the mineralization of organic P. Low molecular organic acids may mobilise various forms of P by: 1) increasing the dissolution of sparingly soluble P minerals, 2) reducing the sorption of P from sorption sites by alteration of the surface characteristics of soil particles (ligand exchange and ligand dissolution), 3) chelating cations (the negatively charged anions of the organic acids compete with the phosphate groups for binding sites on the soil particles, forming strong complexes with aluminium (Al 3+ ) and iron (Fe 3+ ) as well as calcium (Ca 2+ ) ions, leading to a release of phosphate ions into the soil solution) and 4) dissolution of amorphous metal-organic matter coatings that contain occluded P (Attiwill & Adams, 1993; Fox et al., 2011; George et al., 2011). Furthermore, the organic acids may promote the growth of rhizosphere microorganisms that improve plant P acquisition (George et al., 2011). Organic acids have been found to be important for increasing the availability of inorganic as well as organic P (Georg et al., 2011), although relatively little is known about the impact of organic acids on the release of organic P (Fox et al., 2011). Generally, organic acids seem to be particularly important in soils with very low P availability, i.e. very acidic and very calcareous soils (Tyler & Ström, 1995; Ström, 1997; Ström et al., 2002; Palomo et al., 2006) and the efflux is stimulated by P deficient conditions (George et al., 2011). It has been shown that roots of crop plants in sterile hydroponic cultures excrete 30-fold more organic acids under conditions of P deficiency than when well-supplied with P. Citric acid, oxalic acid, malic acid and piscicid acid are among the acids know to solubilise P in soils (Attiwill & Adams, 1993; Lambers et al., 1998). Fox et al. (2011) suggested that organic acids that have log KA1 values greater than 4, increase P release from soils whereas those with log KA1 values less than 4 have little impact on P release. According to Amirbahman et al. (2010), low molecular weight organic acids are more likely to increase the dissolution rate of mineral surfaces than larger molecular weight humic and fulvic acids. This is due to the ability of low molecular organic acids to form a certain type of complex with surface metal centers (mononuclear bidentate complexes) that promote mineral dissolution. Humic and fulvic acids, on the other hand, largely form a kind of complex (monodentate multinuclear complexes) that do not promote mineral dissolution. Most studies with regard to the ability of plants to affect P availability have been performed on annual plant species. However, the ability of trees to increase P availability in soils was well documented in a series of studies in which coniferous trees were planted on pastures. Afforestation of grassland with conifers improved P availability in the topsoil due to enhanced mineralization of organic P, improved solubility of P as a consequence of root and microbial exudates, greater tree root phosphatase activity associated with ectomycorrhizas and favourable soil moisture and temperature regimes (Chen et al., 2008). Several examples of production of organic acids in forest soils are given in Attiwill & Adams (1993) and Fox et al. (2011) gives a number of references for production of phosphatases by tree roots, fungi and bacteria in forest soils. The concentration of low molecular organic acids in forest soils is generally quite low, according to Fox et al. (2011) typically less than 0,01 0,1 mm. 19

22 phosphate) or becomes attached to another phosphate by the energy-rich pyrophosphate bond (e.g. in ATP). Significant amounts of P (up to 50% or more of the total) may be in inorganic form in the foliage of some tree species (Attiwill & Adams, 1993). When the P supply increases from the deficiency to the sufficiency range, the major P fractions in vegetative plant organs also usually increase (Marschner, 2003). In P replete plants, small metabolites, nucleic acids and phospholipids contribute approximately equally to leaf P content (George et al., 2011). According to Marschner (2003), the rates of exchange between various forms of P in the plant are very high. With further increase in supply, only inorganic P as major storage form of P in highly vacuolated tissue increases. However, plants may also store P in two other forms: polyphosphate and phytate (an organic phosphate-containing compund) (Lambers et al., 1998; Marschner, 2003). The storage of P as polyphosphate is widespread among lower plants, but it has also been found in higher plants such as apple leaves (Marschner, 2003). Polyphosphate formation is also evident in hyphae of mycorrhizal fungi, possibly acting as transient energy and storage pools of P in the hyphae (Marschner, 2003). Phytate is the typical storage form of P in grains and seeds (Marschner, 2003) P content in plants Jacobsen et al. (2003) reviewed the nutrient content of a number of coniferous and deciduous tree species located in 115 temperate and boreal forest stands of varying ages in central and northern Europe and Canada. They found that the average P content of eleven beech stands amounted to approximately 22 kg P ha -1 in stem wood, 6 kg P ha -1 in stem bark, 26 kg P ha -1 in branch wood and bark, and almost 6 kg P ha -1 in current twigs and leaves, resulting in a total of around 60 kg P ha -1 in aboveground biomass. The roots contained approximately 19 kg P ha -1 (Jacobsen et al., 2003; Sabine Braun, personal communication). For spruce, the corresponding values were 10, 9, 18 and 20 kg P ha -1 for the aboveground compartments, giving a total of 57 kg P ha -1 in the aboveground biomass (based on 30 stands). The spruce roots contained on average 16 kg P ha -1. In the five oak stands investigated, the stem wood contained approximately 11 kg P ha -1, stem bark 5 kg P ha -1, branch wood and bark 16 kg P ha -1 and leaves 8 kg P ha -1, giving a total in the aboveground biomass of 40 kg P ha -1. Oak roots contained on average 13 kg P ha -1 (Jacobsen et al., 2003; Sabine Braun, personal communication). However, even at very low concentrations, continuous release of organic acids may promote desorption of P and dissolution of amorphous mineral coatings providing P to the soil solution. Fox & Comerford (1992) showed that, over time, the cumulative amount of P released from forest soils was similar when oxalate was added sequentially over time at very low concentrations compared to a single addition of oxalate at higher concentrations. In general, the excretion of organic acids from plants is believed to occur in irregular pulses, thus probably rendering the acids less susceptible to rapid degradation by the soil microbial community, although there have been suggestions that some plant species have a constitutive basal level of exudation (Wouterlood et al., 2004). That rapid uptake and mineralization of the organic acids by the soil microbial community may result in a rapid depletion of the acids, which consequently have a rather short-lived effect on the P solubility in the soil, was shown by for example Van Hees et al. (2003) Leaf uptake There is very little information available regarding leaf uptake of P. However, leaves are capable of acquiring nutrients. For example, volatile N compounds may be taken up through the stomata and nutrients in water on wet leaves are also available for absorption by leaves (Lambers et al., 1998). With regard to P, throughfall data indicate that leaf uptake does occur (see section 4.1.1). At low P availability and/or during the growing season, it may even be relatively large. 6.2 P content of plants and its relation to growth Different forms of P in plants Unlike nitrate (NO 3 - ) and sulphate (SO 4 2- ), phosphate is not reduced in plants but remain in its highest oxidized form (Marschner, 2003). After uptake, it is present as inorganic phosphate, is esterified through a hydroxyl group to a C chain and forms a simple phosphate ester (e.g. sugar Similar values to those of Jacobsen et al. (2003) were found by Nihlgård (1972) and Nihlgård & Lindgren (1977). The P content of trees in three 90-year old European beech (Fagus sylvatica) stands in southern Sweden amounted to kg P ha -1 in stem wood, 3-5 kg P ha -1 in stem bark, kg P ha -1 in branch wood and bark, 5-6 kg P ha -1 in current twigs and leaves, resulting in a total of around kg P ha -1 in aboveground biomass. The below-ground biomass contained kg P ha -1, where 6 kg P ha -1 was in stumps and big roots and 5 kg P ha -1 was in medium roots and small roots, respectively. For a 55-year old spruce stand (Picea abies), the total amount of P in above-ground biomass was 87 kg P ha -1. Stem wood contained 15 kg P ha -1, stem bark 14 kg P ha -1, branch wood and bark 30 kg P ha -1, current twigs and needles 7 kg P ha -1 and older needles 22 kg P ha -1, while stumps and big roots, medium roots and small roots contained approximately 2 kg P ha -1 each, giving a total of around 6 kg P ha -1 for below-ground biomass. Values in the same range have also been found for a number of coniferous and deciduous forest stands in Switzerland (Sabine Braun, personal communication). Somewhat lower values were presented by Zhang & Mitchell (1995), investigating a hardwood stand containing American beech (Fagus grandifolia Ehrh.) and sugar maple (Acer saccharum Marsh.) in the eastern US. In coherence with the European studies, branches contained the largest amount of P (14,3 kg P ha -1 ) in the aboveground biomass, followed by bolewood, bole bark and foliage (7,0, 4,5 and 4,5 kg P ha -1 ) respectively. Fine and coarse roots contained 3,8 kg P ha -1, while herbaceous vegetation represented only 1,0 kg P ha -1 and accumulated litterfall and the forest floor 68,2 kg P ha -1. A large amount of P (3800 g ha -1 ) was translocated by the end of the growing season, conserving P in the ecosystem (Zhang & Mitchell, 1995) P concentrations and nutrient ratios A number of studies have tried to evaluate the critical foliar nutrient concentration of P. Generally, these evaluations are based on empirical relationships between the concentration of the nutrient and leaf symptoms and/or plant growth. Already in the 1970s and 1980s, experimental 20

23 Nutrient treshold values of P (mg g -1 ) and normal P/N ratios in foliage for some common tree species in temperate and boreal forest ecosystems in Europe. Modified from Mellert & Göttlein (2012). Tree species Normal range Deficiency N/P normal range Scots pine (Pinus sylvestris) 1,3-1,9 <1,3 7,4-14,1 Norway spruce (Picea abies) 1,5-2,2 <1,5 6,3-11,7 European beech (Fagus sylvatica) 1,2-1,9 <1,2 10,0-18,9 Oak (Quercus robur, Quercus petrea 1,4-2,1 <1,4 9,3-19,6 research by Ingestad (1979; 1987) emphasized the importance of balanced nutrition for tree growth, i.e. that there is an appropriate relation between the various nutrients essential for tree growth. In some later correlative studies, it was shown that foliage nutrient ratios could be more appropriate than nutrient concentrations to describe the nutrient status of forest tree species in relation to vitality and growth (Cape et al., 1990). In contrast to nutrient concentrations, ratios can reflect possible imbalances and they are less affected by growth dilution, concentrations of soluble carbohydrates and ageing processes (Linder, 1995; Flückiger & Braun, 2003). In particular, this appeared to be valid under conditions such as progressive N saturation of forests (Hüttl, 1990; Mellert et al., 2004). Today, concentrations and ratios are commonly used together. According to optimum nutrient fertilization experiments with forest trees in Sweden, the optimum concentration of P for foliage of Norway spruce is around 1,8 mg g -1 dry weight, or 10% of N (studies compiled by and presented in Linder, 1995). Other studies (see Jonard et al., 2009 and references therein) have suggested the optimum P concentration for trees to be 1,5-2,0 mg g -1 dry weight, while the deficiency level is at 0,8 mg g -1 dry weight. Recently, Mellert & Göttlein (2012) suggested new nutrient treshold values and ratios for the major nutrients based on statistical analysis of the data collected by van den Burg (1985; 1990). The data for P and P/N ratios are presented in the table. The ratios found in nature vary considerably more. According to Güsewell (2004), individual measurements may range from 1 to 100 both within and among species, with an average N/P ratio of terrestrial plants at their natural field sites between 12 and 13. McGroddy et al. (2004) found considerably higher values when investigating the C:N:P stochiometry in forests worldwide. The N:P ratio for temperate broadleaf forest was found to be 28,2 ± 1,5 while it was 21,7 ± 1,7 for conifers (McGroddy et al., 2004). Critical nutrient concentrations and ratios are generally specifically related to growth (de Vries et al., 2007). Mellert & Göttlein (2012) emphasizes, however, that forest trees are faced not only with site-specific nutrient availability but also with other abiotic and biotic site conditions. Adequacy ranges for nutrient concentrations and ratios should thus take other ecological aspects into account, such as cold hardiness, drought resistance and susceptibility of plants to parasite attacks (Flückiger & Braun, 2003). Flückiger & Braun (2003) refer to several papers showing that the susceptibility of tree species to parasite attacks and in- 21

24 festations are increasing if nutrient ratios are unbalanced. Despite that the normal concentration ranges derived from the data of van den Burg (1985; 1990) according to Mellert & Göttlein (2012) are beyond such synecological levels, their ratios agree rather well with those of Flückiger & Braun (2003). As an example, the recommended adequate N/P ratio for Norway spruce is 7-12 in Flückiger & Braun (2003), while in Mellert & Göttlein (2012), it ranges from 6,3 to 11,7. For European beech the corresponding comparison gives values of 10-17,1 (Flückiger & Braun, 2003) and 10-18,9 (Mellert & Göttlein, 2012) Factors affecting the P concentrations The concentration of P in trees varies substantially depending on site, species and environmental conditions. For beech (Fagus sylvatica) growing in southern Sweden, the amount of P was found to be higher in younger leaves and then to decrease throughout the growing season (Tyler & Olsson, 2006). Similar results have been found for coniferous trees, i.e. older needles from previous years commonly have lower P concentrations than current-year needles (Khanna et al., 2007). Niinemets & Tamm (2005) also found that the P concentration decreased continuously in deciduous temperate forest stands, but only after peak physiological activity in midseason. Kutbay et al. (2003), on the other hand, found no differences in P concentration from the time of full leaf expansion to the beginning of senescence for beech trees in a Fagus orientalis forest in the north of Turkey. Leaf nutrient concentrations generally vary with tree age. Older trees have lower foliar macronutrient concentrations than younger ones (Jonard et al., 2009 and references therein). According to Jonard et al. (2009), the decline with age may be attributed to nutrient storage in stems and branches, which decreases the nutrient availability in the soil. In parallel, increasing internal redistribution reduces the nutrient demand of the tree (Kimmins, 1997). Deposition of N may influence the P nutrition of trees in several ways. First, the increased forest productivity as a consequence of increased availability of N may result in dilution effects and lead to nutrient imbalances (Nihlgård 1985; Aber et al., 1989; Mellert et al., 2004). Fertilization experiments with N have resulted in significantly decreased concentrations of P in fertilized trees as compared with control trees under field conditions in stands of Norway spruce (Gundersen, 1998) and in pot experiments with European beech and Norway spruce seedlings (Flückiger & Braun, 1998). In addition, the higher N availability may also reduce the proportion of C allocated to roots (Ericsson et al., 1996; Marschner, 2003) and, subsequently, their ability to take up P. Also climate may influence the P concentration in trees. Drought has been shown to decrease the resorption efficiency of P in trees as well as other plants (Hocking, 1982; Killingbeck et al., 1990; Marchin et al., 2010), suggesting that drought-induced leaf senescence results in net loss of P from a plant. However, Marchin et al. (2010), investigating a mixed hardwood stand, found that tree species which dropped their leaves during drought (Acer rubrum, Liquidambar styraciflua, Liriodendron tulipifera and Nyssa sylvatica) generally resorbed more P from leaves, and, consequently, suffered smaller P losses caused by leaf dessication than species which kept their leaves during drought (among others Fagus grandifolia, Quercus alba, Fraxinus Americana). The strategy of drought-deciduous tree species prevented extensive leaf dessication during drought and successfully averted large nutrient losses caused by leaf dessication. Theoretically, climatic variables could also affect patterns in foliar nutrition by its influence on nutrient transport within the tree (diffusion and mass flow). However, Jonard et al. (2009) found no effect of rainfall on P patterns for broad-leaved trees in France, Walloon and Luxembourg. Fructification, on the other hand, is well-known to be influenced by climate. According to Jonard et al. (2009), a dry warm summer is known to promote seed production the following year (Piovesan & Adams, 2001; Övergaard et al., 2007). Seed production generally requires large amounts of P and translocation from leaves to fruits could occur in masting years, resulting in a decrease in foliar P (Jonard et al., 2009). Harvesting operations have also been found to influence tree nutrition (see section 7.3) P and plant growth As stated in the introduction, P has many essential functions in the plant. Many of the responses of plants to P starvation appear to be initiated, or modulated, by a decrease in the delivery of inorganic P to the shoot and the consequent reduction in inorganic P available for shoot metabolism (George et al., 2011). According to Marschner (2003) and George et al. (2011), this has direct effects on photosynthesis, glycolysis and respiration, resulting in the accumulation of organic acids, starch and sucrose in leaves of P-starved plants (George et al., 2011) and, subsequently, suppression of shoot growth (Marschner, 2003). For details about critical P concentrations and P/N ratios, see previous section. In contrast to shoot growth, root growth is less inhibited under P deficiency and most species partition a greater proportion of their total dry matter into root growth when grown under P deficiency (Marschner, 2003; George et al., 2011). Increased specific root length, increased root hair length and density, root agravitropism or topsoil foraging (producing roots where concentrations of P are relatively large) and the formation of specialised root structures that increase P acquisition are common features. The increased partitioning of carbohydrates to the roots typically result in an increase in the root to shoot ratio of the plant when P is scarce (Marschner, 2003). However, according to Güsewell (2004), the effect of P on biomass allocation is usually less strong than the effect of N. Consequently, plants with a high N/P ratio normally allocate less biomass to roots than plants with the same growth rate but a low N/P ratio. Reduced cytokinin levels in shoots of nutrient-deficient plants are thought to drive the change in biomass allocation by increasing the export of sucrose from shoots to roots (Güsewell, 2004 and references therein). According to Güsewell (2004), the different responses of cytokinins to N and P concentrations in the rooting medium are the likely reason why a high N supply reduces root allocation 12 A sat (umol m -2 s -1 ) y = 58.26x R 2 = Foliage P (%) Photosynthetic rate as a function of foliage P (%) at saturating irradiance and ambient CO 2 in Pinus radiata seedling grown under controlled laboratory conditions. Data and graph provided by Sabine Braun (based on measurements by Bown, 2007). 22

25 even in P deficient plants whereas a high P supply does not reduce allocation in N deficient plants. Under severe P deficiency, however, photosynthesis is also impaired (Marschner, 2003) and there are studies indicating that P may have a strong influence on the photosynthetic capacity of trees (Bown, 2007), especially when P/N levels are low (Reich & Schoettle, 1988). In plants suffering from severe P deficiency, reduction in leaf expansion, leaf surface area and the number of leaves are common. In contrast, the contents of protein and of chlorophyll per unit leaf area are not much affected, often resulting in a darker green colour of P deficient leaves. The photosynthetic efficiency per unit of chlorophyll is, however, much lower in P deficient leaves than in those with an adequate supply of P (Marschner, 2003). The formation of reproductive organs, such as seeds, is usually also retarded by P limitation. According to Güsewell (2004), the low N/P ratio of seeds (values ranging from 1,5 to 15 for wild herbaceous plants) relative to that of shoots suggests that P limitation should affect reproductive output more than growth. Delayed flower initiation and decreased number of flowers as well as premature senescence of leaves are other common features of P deficient plants (Marschner, 2003) Effects of P fertilization on P concentration and growth With regard to P fertilization, a recent review by Fox et al. (2011) showed that P fertilization generally has a positive effect on growth of plantation forests. Volume growth gains ranged from 20% to more than 100% in plantations of several different types of pine and eucalyptus forests following P fertilization near the time of planting. The magnitude of the growth response to P fertilization varied depending on species, soil type, understorey competition and climate (Fox et al., 2011). According to Fox et al. (2011), many of the plantations were grown on P deficient soils, but no data on leaf or tree nutritional status were presented. However, that fertilization with P generally results in improved P nutrition of trees have been shown in several other studies (Polglase et al., 1992a,b; Flückiger & Braun, 1995; Khanna et al., 2007; Prietzel et al., 2008; Prietzel & Stetter, 2010). Flückiger & Braun (1995), for example, found that fertilization with P in combination with other nutrients improved nutritional status, reduced needle loss and increased growth (radial increment, needle length and shoot growth) in a damaged mature Norway spruce stand in Switzerland. Similar results with regard to nutritional status and growth were also found by Prietzel et al. (2008) and Prietzel & Stetter (2010), studying two Scots pine stands in southern Germany. In these two latter studies, the effects of fertilization on the P nutrition was long-term, more than 20 years in one of the stands (75 kg P ha -1 ) and more than 40 years (90 kg P ha -1 ) in the other stand. That fertilization with P generally has long-term effects were also suggested by Fox et al. (2011). According to the studies presented in their review, the response to a single application of 56 kg P ha -1 may last for 20 years or more, and there are several studies where a single application of P in one rotation continued to increase growth in subsequent rotations (Fox et al., 2011 and references therein). These results are supported by those of Trichet et al. (2009), who showed that even low levels of P additions (17-35 kg of P ha -1 ) are sufficient to obtain significant improvements in cumulative growth for Maritime pine in south-western France. Higher rates of application did not yield significantly higher growth responses. Improved P nutrition does not always imply increased growth though. In a Norway spruce stand in south-western Sweden, addition of a vitality fertilizer containing all essential nutrients but N had no impact on the growth of the trees (Nilsson et al., 2001), despite significant increases in the concentration of P in both current-year and older needles as well as branch wood (Nilsson & Wiklund, 1995). Similar results were found by Nilsen & Abrahamsen (2003) in two fertilizer experiments in Norway, where P and N were added to stands of Scots pine and Norway spruce for nine and four years respectively. Also Finzi (2009) reported no overall positive effects of P additions on tree growth in two different types of forest ecosystems (a sugar maple-white ash stand and a red oak-beech-hemlock stand) in the northeastern US. The study was very short and results were recorded only two years following application. However, significant increases were recorded as a response to N addition. The results of P fertilization studies in arctic and alpine ecosystems have been mixed, with some plant communities (such as alpine wet meadow) showing signs of P limitation to primary productivity, whereas other communities are either N-limited or co-limited by N and P (Weintraub, 2011). 6.3 Retranslocation of P McGroddy et al. (2004), investigating the C:N:P stochiometry in forests worldwide, found that the C:P ratios in litter were consistently higher than in comparable foliar data sets, suggesting that resorption of P is a globally important mechanism. Low concentrations of P in litter and efficient resorption of P during leaf senescence was also concluded by Güsewell (2004) as important conservation mechanisms for P. However, the fraction of mineral P withdrawn from leaves before their abscission varies considerably. Hagen-Thorn et al. (2006) reported rates between 37 and 59% for four common deciduous tree species in Lithuania. Kutbay et al. (2003) found that resorption rates for various plants in a Fagus orientalis stand in northern Turkey varied between 50 and 80%. The rate did not differ significantly between species. When compiling foliar P resorption rates for some other deciduous and evergreen species, Kutbay et al. (2003) found values ranging from 0 to 80%. In a review comparing over a hundred deciduous and evergreen shrubs and trees, Aerts (1996) reported that 52% of the P was withdrawn from the leaves. According to Aerts (1996), resorption efficiencies are not clearly related to the fertility of sites. Nutrient resorption did decrease upon enhanced nutrient supply in 35% of the cases. However, in 57% of the cases, there was no response. Evergreen shrubs and trees showed especially low responsiveness to soil fertility. Other studies have indicated that the efficiency of nutrient withdrawal and retranslocation are completely unaffected by soil fertility (Chapin & Moilanen, 1991; Helmisaari, 1992; Bown, 2007), but are controlled by internal factors, primarily growth rate (Nambiar & Fife, 1987; 1991). Resorption rates have also been suggested to vary with leaf life span, leaf area span, overall leaf nutrient content, plant functional type and environmental conditions such as drought (Del Arco et al., 1991; Kimmins, 1997; Niinemets & Tamm, 2005; Hagen-Thorn et al., 2006). Studies of arctic and alpine ecosystems have shown that retranslocated and stored P may be of great importance for early growth in cold climates (Weintraub, 2011). 6.4 P in litterfall The amount of P returned to the soil by litterfall is primarily a function of biomass, plant P concentration and the extent of retranslocation which depend on among other factors climate, soil type, species and age of the forest (see sections 6.2 and 6.3). Fox et al. (2011) found that P content in the forest floor varies from less than 10 kg ha -1 to more than 300 kg ha -1. Laiho & Prescott (2004), reviewing the decay and nutrient dynamics of coarse woody debris in northern coniferous forests, presented a compilation of litter inputs and litter nutrient contents from a number of studies on several different coniferous tree species. Input of P was shown to range from 0,001 to 0,06 g m-2 yr -1 in boles, while the content in fine woody debris (twigs, branches, bark fragments and reproductive tissue) constituted between 0,024 and 0,05 g m -2 yr -1. Input from needles was considerably higher, between 0,059 and 0,38 g m -2 yr -1. Coarse woody debris constituted between 1 and 12% of the total input, and Laiho & Prescott (2004) concluded that nutrient input from coarse woody debris are considerably smaller than those in other litter types, because of the low nutrient concentrations in coarse woody debris. Consequently, coarse woody debris is generally a sink for P initially, but becomes a source in later decay (Laiho & Prescott, 2004). The P content in the forest floor was found to range from 2 to 22 g m -2, while the upper 50 centimeters of the soil had values ranging from 9 to 518 g m -2 (Laiho & Prescott, 2004). In a hardwood forest in the eastern US, litterfall returned 3100 g P ha -1 yr -1 to the forest floor (Zhang & Mitchell, 1995). The importance of root litter for DOP in soils was emphasized by Uselmann et al. (2009). Root litter may, under certain circumstances, even contribute more to DOP than leaf litter (Uselman et al., 2009). According to Fox et al. (2011), fertilization often increases the growth rate of forests and, consequently, litterfall. Furthermore, P fertilization has been shown to increase the P content in litter and may increase the return of total P in litterfall between 150 and 400% compared with unfertilized forests (Fox et al., 2011). 23

26 7. P OUTPUT The principal pathways by which P is lost from an ecosystem are: 1) erosion of P carrying soil particles, 2) P dissolved in surface runoff water and 3) plant removal (Brady & Weil, 1999). It is a well-known fact that soils used for agricultural production loose substantial amounts of P to streams every year, partly because of the P enrichment of agricultural soils, but also as a consequence of agricultural operations such as tillage (Brady & Weil, 1999). In natural terrestrial ecosystems where soils are covered by undisturbed forests or natural grasslands, on the other hand, P is relatively closely cycled between soils and biota. The absence of a gaseous phase, the low solubility of inorganic P and the limited soil erosion under good vegetation cover are key determinants of this close cycling (Tiessen et al., 2011). According to Uselman et al. (2009), a large proportion of the P lost from forest ecosystems is in the organic form. 7.1 P erosion Since erosion tends to transport predominantely clay and organic matter fractions of the soil (which are relatively rich in P), eroded sediment is often substantially enriched in P (Brady & Weil, 1999; Thiessen et al., 2011). Pimentel (2006), reviewing the factors causing soil erosion, concluded that soil erosion rates depend on past and present land use, soil type, climate and land-surface forms. Under good vegetation cover, soil erosion, and consequently also P loss, is generally limited (Pimentel, 2006; Tiessen et al., 2011). This is a consequence of raindrop and wind energy being dissipated by the biomass layer and the topsoil being held by the biomass (Pimentel, 2006). However, disturbances such as timber harvest or wildfires, may increase the loss of P substantially, primarily via eroded sediment (Brady & Weil, 1999). In a steeply sloped sandy loam site in North-western Spain, the P 24 loss of wildfire-affected plots (9,1 kg ha-1) was more than 6-fold that of control plots (1,4 kg ha-1) in the year following the fire (Brady & Weil, 1999). Moderately burned plots (4,3 kg ha-1) showed P losses in between those of control and wildfire. Brady & Weil (1999) emphasizes, however, that levels of P lost from humid-region forests on soils low in available P are usually smaller. According to Pimentel (2006), a minimum of 60% cover is necessary in forested areas to prevent serious soil erosion and land slides. With regard to soil type, soils with medium to fine texture, low organic matter content and weak structural development are generally most easily eroded (Pimentel, 2006). Typically, these soils have low water infiltration rates and are therefore subject to high rates of water erosion and the soil particles are also easily displaced by wind. Furthermore, the topography of a landscape, and its rainfall and/or wind exposure all combine to influence the susceptibility to erosion (Pimentel, 2006). In general, erosion of P carrying soil particles may vary between 0,1 and 10 kg ha-1 annually on organic and mineral particles, the higher figure most likely apply only to cultivated soils (Brady & Weil, 1999). 7.2 Leaching of P Because soluble inorganic forms of P are strongly adsorbed by mineral surfaces, the loss of P through leaching is generally very low (Attiwill & Adams, 1993; Brady & Weil, 1999). The exception is organic soils or sandy soils to which very high amounts of animal manure have been added (Brady & Weil, 1999). According to Brady & Weil (1999), losses through leaching may amount to between 0,01 and 3,0 kg ha-1 annually, the higher figure most likely apply only to cultivated soils. Newman

27 (1995), reviewing P input to terrestrial ecosystems, gave some examples of studies that had measured P leaching. In a tropical moist forest, the amount of P leached was in the range of 0,01-0,7 kg ha -1 yr -1 and for an unfertilized permanent grassland it was 0,4 kg ha -1 yr -1. Zhang & Mitchell (1995) found that leaching of total P from a hardwood stand in eastern US amounted to 6,5 g ha -1 yr -1, constituting 26% of the input by precipitation. Löfgren et al. (2009) presented values of 0,06 to 0,07 kg P ha -1 yr -1 for an approximately 100-year old forest stand in northern Sweden, mainly consisting of Norway spruce and Scots pine. Hitherto, there are no well established controls of P leaching from forested areas. Kopácek et al. (2011a) found that P leaching from soils in catchments along an elevation gradient in the Tatra mountains and Bohemian forest in the Czech republic was positively related to DOC leaching. Similar close co-variances between lake water concentrations of total P and organic C and N was observed for numerous European mountain lake districts (Camarero et al., 2009), something that according to Kopácek et al. (2011a) indicate that their terrestrial co-export from catchment soils may be a general feature. That P export was closely related to TOC export was also found by Kortelainen et al. (2006). Also dissolved inorganic C (DIC) has been found to influence the mobilisation of P (Amirbahman et al., 2010). Although a poor correlation between DIC and P, DIC clearly altered P mobilization in soil column transport experiments. Ligand exchange was given as the most likely mechanism behind the increased mobilization (Amirbahman et al., 2010). The associations between DOC and DIC and P indicates that controls for DOC and DIC may also be valid for DOP (Kopácek et al., 2011a). However, this remains to be thoroughly tested. Kaiser et al. (2000), investigating the soil water of one Scots pine stand and one European beech stand in Germany, found, for example, that DOP was mainly concentrated in the hydrophilic DOM fraction which was more mobile in the mineral soil than the hydrophobic one. Consequently, the concentrations and fluxes of DOP decreased less with depth than those of DOC. That controls that have been established for DOM are valid for DOP was also questioned by Kalbitz et al. (2000). On a European scale, Mattsson et al. (2009) found that concentrations and loads of DOP were positively associated with percentage of urban and agricultural areas, population density and the number of days with precipitation over 10 mm. In a similar study of boreal catchments, Mattsson et al. (2005) found that the proportion of agricultural land and upstream lakes explained 50-88% of the variation in the export of different forms of P (and N). Kopácek et al. (2011a) found that lake water DOC:TP concentrations increased with catchment soil cover. Furthermore, leaching was small in the mountaineous area while considerably higher in the forested catchments. According to Kopácek et al. (2011a), this difference is likely to be due to differences in soil and vegetation pools, which are smaller in the more mountaineous area. A negative association between slope and total P export was found by Kortelainen et al. (2006). However, the total stem volume within the stand was not found to be an important predictor for export of total P (Kortelainen et al., 2006). 7.3 Effects of tree harvest and forest management on P The major loss of P as a consequence of forest management operations is the direct removal of P in tree parts. Accumulation of P in stems of trees and in entire trees has been found to be highly variable (Mann et al., 1988). According to Brady & Weil (1999), plant removal may remove 5 to 50 kg ha -1 annually of P from an ecosystem, depending on plant species, soil type and cultivation system. A study of eleven forest stands located throughout the US showed that the harvest of conifer logs (sawlog or pulpwood harvest) removed between 5 and 56 kg P ha -1 at clear-cut (only direct removal of nutrients in harvested wood), whereas whole-tree harvest removed between 10 and 96 kg P ha -1 (Mann et al., 1988). Corresponding values for hardwoods were 4 to 41 kg P ha -1 for logs and 19 to 47 kg P ha -1 for whole-tree harvest. These results are supported by calculations based on data provided by Jacobsen et al. (2003), showing that whole-tree harvest may remove up to five times as much P as conventional harvest (if the bark is left on site; Sabine Braun, personal communication). Looking at net nutrient fluxes (i.e. deposition, leaching and harvest removals), Mann et al. (1988) found that approximately 70% of all stands showed a net loss of P when sawlogs were harvested, while 80% showed a net loss when whole trees were harvested. In accordance with the numbers presented by Mann et al. (1988), Federer et al. (1989), also investigating a number of forest stands (both conifer and hardwood) in the US, reported a removal of between 20 and 50 kg P ha -1 yr -1 when wholetree harvest was applied at clearcut. According to Federer et al. (1989), whole-tree harvest removed close to 90% of the above-ground pool of P, but only less than 3% of the total pool of P at all sites. With regard to indirect effects of forest management on the P balance of forest ecosystems, recent model simulations for the export of P from Sweden to the Baltic sea, the Kattegatt and the Skagerrak indicate that 27% of the net nutrient load of P originate from forest land (Brandt et al., 2008). However, only 2% of this load was attributed to clear-cutting, the rest was assumed to be governed by natural processes. Although small on a national scale, disturbances such as forestry operations may have significant impacts on nutrient exports on a local scale, detrimentally affecting surface waters of small catchments. In Sweden, a study by Löfgren et al. (2009) showed that both runoff, transport of total P and suspended solids were enhanced in the harvested catchment as compared with the reference catchment. Uggla & Westling (2003) also found indications that the concentrations of total P may increase directly after harvest and Mann et al. (1988) reported that hydrological losses of P was slightly higher after whole-tree harvest compared with sawlog or pulpwood harvest in a study of eleven forest stands located throughout the US. Federer et al. (1989), on the other hand, found no harvest-induced leaching of P after whole-tree harvest when investigating a number of conifer and hardwood sites in the US. The reasons for the increased indirect P (and other nutrients) export at harvest probably include increased runoff as a consequence of reduction in evapotranspiration after harvest, increased erosion and sediment transport as a result of harvesting activities (i.e. logging tracks) and site preparation and increased mineralization due to increased insolation (Löfgren et al., 2009). Binkley et al. (1999) synthesized the information available from published studies about the effects of forest fertilization on water quality. They found that fertilization with phosphate can lead to increased peak concentrations of >1 mg l -1, but that annual averages usually remained <0,25 mg l -1. Fertilization may thus increase the average P concentrations several-fold, but according to Binkely et al. (1999), the transient timing of increases in P concentrations together with P removal and dilution downstream probably results in small overall effects on the aquatic ecosystems. 25

28 8. A CONCEPTUAL MODEL 8.1 Causal loop diagram a tool to visualize conceptual models Causal loop diagram (CLD) is a tool that is commonly used to construct conceptual models. As indicated by their name, CLDs are diagrams that map the causal relationships between the parameters constituting a system, thus creating an integrated network of relationships that is presented in a readable and transparent way. The arrows in the CLD represent the causalities as they carry the effect of the variable at the tail of the arrow (origin) to the variable at the head of the arrow (target). As CLDs are conceptual, the only possible changes in the variables are either an increase or a decrease. All changes take place over time, but are driven by changes in another variable. A state of no change is also possible, and thus a variable may be constant over time. An arrow with a positive causality sign implies that a change in the origin variable drives a similar change in the target variable (i.e. an increase in the origin leads to an increase in the target, and a decrease leads to a decrease). Arrows with negative causality signs imply a change in the opposite direction (i.e. an increase in the origin leads to a decrease in the target, and a decrease leads to an increase). This means that the arrows themselves do not indicate whether a variable increases or decreases with time. Bearing in mind this simple rule regarding arrows, CLDs are able to reproduce dynamic behaviour that involves changes between increases and decreases in different variables as the entire system changes with time. In a balancing system, feedback loops move the system in the direction towards fluctuation around an equilibrium point. A feedback loop may also be reinforcing (i.e. move the system away from the equilibrium point). 8.2 A CLD of P cycling in temperate and boreal forest ecosystems To visualize how various environmental factors may influence the P cycling in temperate and boreal forest soils, a CLD of P cycling was constructed. The CLD represents the processes governing the P cycle in a temperate or boreal forest stand and is based on the information given in sections four to seven. Below, a description of the diagram is given. For information supporting the links provided in the diagram, please refer to the sections mentioned above. The P that is deposited to forest ecosystems may be taken up directly by the leaves in the canopy, or enter the soil pool of P. If plant P is low, the P uptake through leaves will increase, and, consequently, the amount that is leached from the canopy and enter the soil via throughfall will decrease. If plant P is high, on the other hand, the canopy uptake of P will decrease, and more P will enter the soil via throughfall. Another factor that influences how much P that is transferred to the soil from the plant is litterfall. In general, the more P that the plant contains, the more P is returned to soil as litter. Conversely, if the plant contains very little P, more P will be retranslocated within the tree and less P will be returned to soil through litterfall. Litterfall P enters the soil in the form of organic P and thus contributes to the soil organic P pool. If litterfall P decreases, the pool of soil organic P decreases. If there is little P in relation to C, the C/P ratio of the soil increases. If the C/P ratio increases, P immobilization increases with a subsequent increase in soil organic P. Conversely, if soil C/P decreases, immobilization decreases. If immobilization decreases, the soil organic pool of P decreases. If litterfall P increases, on the other hand, the pool of soil organic P increases. If soil organic P increases, mineralization increases. Since mineralization contributes P to the pool of bioavailable P in solution, a higher mineralization rate implies a higher amount of bioavailable P. The plant may control the mineralization process by exuding phosphatases that influence the mineralization rate. If plant P decreases, the exudation of phosphatases increases, and the mineralization rate increases. The bioavailable pool of P in solution is also controlled by how much soil inorganic P that is adsorbed to the exchange sites of clays, precipitated as Al-, Fe- and Ca- compounds and weathered. If adsorption and net mineral precipitation decrease, and weathering increases, bioavailable P increases. If adsorption and net mineral precipitation instead increases, and weathering decreases, bioavailable P decreases. If net mineral precipitation increases, the amount of soil inorganic P bound to Ca, Al and/or Fe increases. If adsorption increases, the amount of clay-bound P increases. If bioavailable P decreases, leaching decreases. Plant P depends on root uptake and canopy uptake of P. If they increase, the P in the plant increases. If plant P increases, P is available for all the processes for which it is needed, and the shoot growth may thus increase. If shoot growth and plant P increases, photosynthesis increases with a subsequent increase in the production of non-structural carbohydrates. An increase in the amount of non-structural carbohydrates increases the amount of shoot carbohydrates available for shoot growth, with a subsequent increase in shoot growth. If shoot growth is limited for some reason (for example due to lack of a nutrient), less shoot carbohydrates will be utilized, and more carbohydrates will then be transported to the roots and utilized for root growth (including mycorrhizal associations) or production of organic acids. The more root growth, the higher the root uptake of P and the smaller the amount of bioavailable P left. For a more thorough description of the C allocation described above, see Jönsson (2004; 2006). Organic acids may influence P availability in several different ways. First, they lead to a decrease in the amount of reactive Ca, Al and Fe in the soil solution, and, consequently, to a decrease in net mineral precipitation, and, consequently, a decrease in inorganic P bound to Ca, Al and Fe. Second, higher concentrations of organic acids imply a decrease in the net adsorption of P, and, consequently, a decrease in clay-bound P. Seeds generally contain high amounts of P. If frutification increases, more P will be utilized and less P will be left in the plant pool. Similarly, the greater the harvest, the smaller the amounts of P left in the plant pool. Another way for P to leave the system is through erosion. If root and shoot growth increases, erosion will decrease. If the growth decreases, on the other hand, erosion will increase, resulting in a decrease in claybound P and soil organic P in the system. 26

29 The conceptual model of P cycling in temperate and boreal forest ecosystems. For information on how to interpret the diagram, see sections 8.1 and

30 9. THE PROTOTYPE MODEL The plant compartment of the prototype model connects to the soil compartment through throughfall and uptake, and to the litter compartment through litterfall. 28

31 A prototype computer model was built to describe the P cycle in a boreal or temperate forest stand. The prototype was built using the computer modelling tool STELLA (V9.7.4, isee systems, 2010). The prototype is a functional representation of the P cycle as described in the review, but with limited boundaries and a simplified scope. The prototype simulates the flow of P between different ecosystem compartments: biomass, soil organic and mineral pools and soil solution. The prototype as presented in this report is not aimed as a tool for scenario analysis, but rather as a first test of the hypotheses describing the different processes governing the P cycle. It is however fully functional as a tool for sensitivity analysis and for verifying the viability of different hypotheses in terms of conservation of mass balances and relative sizes of P pools and fluxes. The prototype provides the basis for constructing a fully integrated P module with a larger model platform, as well as a being a means for testing different solutions to describe different processes numerically. The model is based on the processes summarized in the CLD (see section 8), and is parameterized based on the information presented in the review. For simplification, the prototype was built into three interlinked sectors: 1) the plant sector, 2) the litter and humus sector and 3) the soil sector. The plant sector simulates the flow and the storage of C and P in plants. According to the architecture adopted in ForSAFE, plants are assumed to consist of foliage, wood and roots. Unlike in ForSAFE, where photosynthesis is physiologically modelled, photosynthesis is assumed in the prototype to fulfill a hypothetical growth curve, and is amended by foliage P concentration. Carbohydrates, the products of photosynthesis, are subsequently allocated to the different parts of the plant according to hypothetical ratios, which in ForSAFE are dynamically modelled. C allocation can be limited by P availability, as the plant is restricted to fulfill the minimum tissue P concentrations given as parameters. Respiration is not simulated in the prototype, but is included in ForSAFE. P is allocated to the three plant compartments from a labile pool assumed to be a Langmuir function of plant size according to the hypothesis used in ForSAFE. The deficit in this labile pool define potential P uptake, referred to as sink strength. P retranslocation prior to litterfall is assumed to occur only in foliage and roots, and not in wood. Soil organic matter is divided into litter, humus and microbial biomass. In ForSAFE, soil organic C is divided between four pools with different turnover rates. The turnover rates in the prototype are hypothetical, The litter and humus compartment of the prototype model receives litterfall from the plant compartment and sends net mineralisation to the soil compartment. 29