Calcium Watershed Addition

Proposal

Project Descriptions

Correspondence

I. Introduction

pH is a "master variable" which influences the structure and function of ecosystems through effects on microbial processes (Adams et al. 1978; Zelles et al. 1987, Persson 1988), abiotic reactions in soil and water (Reuss and Johnson 1986; Nodvin et al. 1988), the bioavailability of nutrients and toxic substances (Reuss and Johnson 1986; Cronan and Grigal 1995), and the distribution and abundance of biological species (Adams et al. 1978; Persson 1988). pH, in turn, is influenced by biogeochemical processes involving the transfer of major ionic solutes (Driscoll and Likens 1982; van Breemen et al. 1983; Binkley and Richter 1987). In forest ecosystems, these element transfers include atmospheric deposition, uptake of ionic nutrients by vegetation, mineral weathering, soil chemical processes (e.g., cation exchange, secondary mineral formation, anion adsorption), mineralization of soil organic matter, microbial transformations involving ionic solutes, and gaseous and drainage losses of major elements.

Soil development and acidification are strongly altered by the supply of basic cations (Ca²⁺, Mg²⁺, K⁺ Na⁺), primarily Ca²⁺, from soil minerals (van Breemen et al. 1983, Binkley and Richter 1987). In areas where weathering release of basic cations is high, basic cation supply may be sufficient to neutralize inputs of organic and carbonic acids from mineralization processes, and to provide nutrient cations for biomass accumulation and leaching while maintaining a relatively high soil base saturation (%BS). At lower weathering rates, insufficient supply of basic cations may decrease soil and drainage water pH, and lower %BS in soil. Through these interactions, basic cation supply exerts an important influence on the structure and function of a forest ecosystem. Base-rich forests differ compositionally and functionally from base-poor forests. Notably, increased supply of basic cations and soil pH: i) alters the composition of the soil decomposer community and increases microbial activity (Zelles et al. 1987; Illmer and Schinner 1991); ii) increases mineralization of soil organic C and alters nitrogen (N) mineralization (Adams et al. 1978; Marschner et al. 1989), and increases nitrification (Tamm and Popovic 1989); iii) changes the species distribution and abundance of soil invertebrates (Persson 1988); iv) affects root growth, plant nutrient uptake and plant productivity (Smallidge et al. 1993); and v) regulates the acid-base status of stream drainage and the composition of stream biotic communities (Baker and Christensen 1991).

Areas exhibiting low Ca²⁺ supply include many of the most acid-sensitive watersheds in the northeastern U.S. (Landers et al. 1988). These watersheds are generally characterized by shallow deposits of surficial materials, soil minerals with slow rates of chemical weathering, and low concentrations and pools of exchangeable basic cations in soil (April and Newton 1985; Driscoll 1991; Eilers and Selle 1991). The geochemical consequences of acidic deposition to forested watersheds include depletion of labile pools of nutrient cations (Kirchner 1992; Bailey et al. 1996) and the mobilization of elevated concentrations of Al from soil to drainage waters (Cronan and Schofield 1979, 1990).

Our long-term studies at the Hubbard Brook Experimental Forest (HBEF), New Hampshire have shown marked changes in the acid-base status of soil and drainage waters over the last 30 years resulting from changes in atmospheric deposition (Figures 1; Likens et al. 1996). Decreases in atmospheric deposition of sulfate (SO₄^2-), coinciding with decreases in sulfur dioxide (SO₂) emissions in the eastern U.S. (Butler and Likens 1991), have resulted in marked decreases in streamwater concentrations of SO₄^2-. Despite these reductions, there has been only very limited increases in stream pH. Rather, we have observed stoichiometric decreases in the sum of basic cation concentrations (C_B) in streamwater. Budget calculations suggest that atmospheric inputs of strong acids to the HBEF have resulted in marked depletion of labile pools of Ca²⁺ from the forest ecosystem (Figure 1; Likens et al. 1996). This process has delayed the recovery of drainage waters from decreases in atmospheric deposition. Coincidently, we have observed a number of long-term changes in the abundance or inorganic nutrition of biotic populations, including sugar maple (Acer saccharum; Likens et al. 1997); acidophilic and calciphilic herbs (Siccama, unpublished; see below); gastropods (Hamburg and Strayer, unpublished), and stream invertebrates and amphibians (Likens, unpublished). Our observations are consistent with other studies from eastern North America and Europe which have shown limited recovery of surface waters from acidic deposition, depletion of soil pools of basic cations and, changes in the abundance of acidophilic and calciphilic species (Dillon et al. 1987; Kirchner 1992; Graveland et al. 1994; Falkengren-Grerup 1995).

In this study, we propose to investigate the role of Ca²⁺ in regulating the structure and function of a northern hardwood forest. This will be accomplished by the experimental addition of wollastonite (CaSiO₃), a

Figure 1. Annual fluxes of Ca²⁺ for w6 of the HBEF during 1940-1963 (estimated) and 1963-1994 (measured) P (o) is bulk precipitation, W (D) is weathering release, S (l) is streamwater loss, B (¨) is net biomass storage and D(-) is net release from labile soil pools (exchangeable + organically bound), obtained by difference. Data during 1955-56 are from Junge and Werby (1958) (after Likens et al. 1996).

readily weatherable Ca-silicate mineral, to watershed 1 (w1) at the HBEF. The dissolution of wollasonite will supply Ca²⁺, H₄SiO₄ and acid neutralizing capacity (ANC) to solution:

CaSiO₃ + H₂O + 2H⁺ = Ca²⁺ + H₄SiO₄

We suggest that supplying Ca²⁺ by the dissolution of a silicate mineral that is similar to the natural weathering source of this element in forest ecosystems throughout the Northeast is a much more experimentally sound treatment than adding CaCO₃, as has been done in several previous studies (Adams and Dickson 1973; Nihlgard et al. 1988). Moreover, the wollastonite that we will use in this study has a distinct ⁸⁷Sr/⁸⁶Sr ratio compared to other sources at Hubbard Brook (e.g., precipitation, weatherable minerals). This stable isotopic tracer will enable us to trace the fate of the added Ca²⁺ in the ecosystem with exceptionally high sensitivity.

We envision the proposed experiment as a long-term study, with experimental measurements planned for ca. 50 yr. The Ca²⁺ addition study will be conducted in two phases. In the first phase, 850 kg/ha wollastonite will be added to w1 by helicopter to increase the %BS of the soil from 10% to approximately 19 %. This change reflects our estimate of the Ca²⁺ depleted from soil at the HBEF over the last 50 yr due to atmospheric deposition of strong acids (Likens et al. 1996). We will study the ecosystem response to this experimental treatment by measuring: throughfall; litter mass and chemistry; biomass, species composition and chemistry of herbaceous vegetation; biomass, growth, species diversity and chemistry of overstory vegetation; soil and soilwater chemistry; microbial activity; composition of soil invertebrates; stream chemistry and aquatic biota. During this first phase of study (0-10 yr), we will evaluate the ecosystem response to the return of the Ca²⁺ that was depleted from labile soil pools due to atmospheric deposition of strong acids. In the second phase (10-50 yr) , we will add approximately 660 kg/ha wollastonite to further increase the % BS of the soil to 25%. ( The final dose for the second phase application will be refined based on the results obtained from the first addition.) This longer-term experiment will enable us to test more fundamental hypotheses of the role of Ca²⁺ supply in regulating the structure and function of the northern hardwood forest. The experimental treatments will be compared to our long-term ecosystem measurements in and around watershed 6 (w6), the biogeochemical reference watershed.

Through this proposal, we seek funds for five years in which we will: i) make premanipulation measurements and install the infrastructure necessary to conduct the manipulation; ii) obtain and apply wollastonite to w1; iii) carry out fundamental measurements of biogeochemical, phytosociological and biotic change in response to treatment ; and iv) conduct intensive measurement of ⁸⁷Sr/⁸⁶Sr ratios in response to the experimental manipulation. At the end of the five-year study period, measurements of the longer-term response of the ecosystem to the experimental treatment will be included as part of the Hubbard Brook Long-Term Ecological Research (LTER) project. Note that if this project is funded, we will encourage the submission of "satellite" proposals by interested researchers to study aspects of the ecosystem response to Ca²⁺ addition in more detail.

II. Objectives and Hypotheses

The role of soil base status as a driver of the structure and function of forest ecosystems is difficult to evaluate using comparative studies because of coincident differences in other environmental and biotic factors between base-rich and base-poor ecosystems, and because of the complex biogeochemical and ecological mechanisms and interactions that govern this role. The long-term objective of the proposed research is to evaluate the mechanisms whereby Ca²⁺ availability regulates the structure, function and composition of northern hardwood forests. This long-term objective would be attained through a comprehensive suite of measurements of a base-enriched watershed ecosystem (w1) in comparison with our reference watershed (w6) at the HBEF. In the initial stage of the proposed research we would add sufficient Ca²⁺ to the soil exchange complex to replace the Ca²⁺ which was depleted from labile pools over the past several decades by the atmospheric deposition of strong acids. In this phase, we will evaluate the short-term (10 yr) response of forest and aquatic ecosystems to the increase in Ca²⁺ availability and consequent changes in soil base saturation, pH and Al availability. In particular, we will test the following hypotheses concerning the direct effects of Ca²⁺ addition. We hypothesize that the experimental addition of wollastonite to w1 will:

III. Rationale and Justification

Calcium is a macronutrient which is required for several physiological processes, including the formation of cell walls and cell membranes, lipid synthesis and mitosis (Devlin and Witham 1983). At low environmental concentrations Ca²⁺ may be a limiting nutrient for a variety of terrestrial and aquatic organisms, including plants (Timoney et al. 1993), macrofungi (Ivanov 1987), birds (Graveland and Drent 1997), rodents (Hansson 1990), gastropods (Graveland et al. 1994), soil and aquatic invertebrates (Krueger and Waters 1983; Hagvar 1990) and diatoms (Planas 1996). However, the ecological effects of Ca²⁺ are closely linked to related responses of pH and Ali. A critical consequence of low Ca²⁺ supply and low pH conditions is the mobilization of Al from soil (Cronan and Schofield 1979, Cronan 1994). Elevated concentrations of Ali are evident in soil solutions of northern forests with low % BS (<15%) and high concentrations of strong acid anions (Cronan and Schofield 1990). Plants exhibit deleterious effects following exposure to ionic Al either through: i) antagonistic interference of cation assimilation or ii) irreversible damage of cells due to interactions with susceptible biomolecules (Sucoff et al. 1990, Cronan and Grigal 1995). In soils with low % BS , the chemistry of Ca²⁺, Ali and H⁺ are closely coupled through soil exchange reactions (Reuss and Johnson 1986). Watersheds which are characterized by low Ca²⁺ supply and experience elevated concentrations of strong acid anions inevitably exhibit high concentrations of H⁺ and Ali in drainage water (Driscoll and Newton 1985). Because of this close linkage, it is difficult to separate the response of organisms to low Ca²⁺ concentrations and high concentrations of H⁺ and Ali.

This problem of coincident changes in Ca²⁺, pH and Al led Cronan and Grigal (1995) to conclude that Ca:Al ratios were a useful indicator of stress associated with acidic deposition and harvesting in forest ecosystems. Based on laboratory and field experiments involving several species, Cronan and Grigal (1995) concluded that when Ca:Al molar ratios were < 1 in soil water, there was a 50% risk of deleterious effects on tree growth and nutrition. This risk increased to nearly 100% when the Ca:Al molar ratio was < 0.2. They also suggested several factors which could serve as indicators of forest stress, including: i) % BS < 15% of effective cation exchange capacity (CEC), ii) the presence of elevated concentrations of strong acid anions, iii) Ca:Al molar ratios < 1.0 in soil solutions, iv) Ca:Al molar ratios of < 0.2 in fine roots and v) Ca:Al molar ratios < 12.5 in foliar tissue. Based on long-term (12 yr) soil solution measurements at the HBEF, the mean molar Ca:Ali ratio for high elevation (750 m) spruce-fir-birch stands was 1.27 in Oa soil solutions, decreasing to 0.40 in Bh and 0.27 in Bs soil solutions. In high elevation (730 m) deciduous stands, the mean Ca:Ali molar ratio was 1.21 in Oa horizon solutions, decreasing to 0.59 in Bh solutions and 0.39 in Bs horizon solutions (Likens et al. 1996).

We recently summarized the biogeochemistry of Ca²⁺in the northern hardwood forest at Hubbard Brook (Figure 2; Likens et al. 1997). One of the most distinguishing characteristics of the Ca²⁺ cycle at the HBEF is the marked depletion of labile Ca²⁺ pools (see Figure 1). Using ⁸⁷Sr/⁸⁶Sr measurements, Bailey et al. (1996) estimated similar rates of soil Ca²⁺ depletion at nearby Cone Pond Watershed. It appears that the depletion of exchangeable Ca²⁺ from soil has contributed to the delay in the recovery of drainage water ANC from acidic deposition (Likens et al. 1996).

The processing of acids and bases in forested watersheds varies markedly with landscape position (Likens et al. 1994; 1997). Hence, a critical issue of practical and conceptual importance is the spatial variability of biogeochemical processes across the watershed landscape. At the HBEF we have assessed this variability in the reference watershed (w6), by considering three elevational zones: i) the spruce-fir-white birch (SFWB) zone which occurs over the range 750-790 m; ii) the high elevation hardwood zone (HH; 650-750 m); and iii) the low elevation hardwood zone (LH; 550-650 m). Soils near the ridges of the HBEF are thin, with areas of exposed bedrock and boulders. With decreasing elevation (and increasing drainage area), there is increasing depth of soil and glacial till. With increasing soil depth, the CEC in the soil and pool of weatherable minerals increases. This spatial pattern in the supply and retention of Ca²⁺ across the landscape is a major factor regulating the elevational variation in the acid-base status and associated ecosystem characteristics of forest watersheds at the HBEF. For example, the headwaters of the SFWB zone exhibit chronically acidic conditions (ANC -96 meq//L) associated with very low concentrations of Ca²⁺ (14 mmol/L). With increasing drainage area and soil depth at lower elevations, ANC values increase, approaching positive values at the weir (-5 meq/L; Ca²⁺ 21 mmol/L) and reaching positive values in the main Hubbard Brook at the base of the Hubbard Brook Valley (ANC 55 meq/L; Ca²⁺ 35 mmol/L: Likens et al. 1997).

The effects of this spatial pattern on the supply of Ca²⁺ are also evident as pools and concentrations of exchangeable Ca²⁺ are low at high elevations within the experimental watersheds and increase with decreasing elevation (Likens et al. 1997). Additionally, low concentrations of Ca²⁺ in drainage water and on the soil exchange complex coincide with low concentrations of Ca²⁺ in foliage of sugar maple and yellow birch (Betula allegheniesis) and smaller pools of Ca²⁺ in living biomass at high elevation. Vegetation at high elevation has low Ca²⁺ content in litter and thus recycles little Ca²⁺ to the forest floor. This biotic feedback under infertile conditions exacerbates the low Ca²⁺ availability within the SFWB zone.

The high elevation zones within the experimental watersheds are more susceptible to acidic deposition than the more base-rich, lower elevation zones. Consequently, concentrations of Ali are elevated at high elevation (Johnson et al. 1981; Lawrence et al. 1986). Chronically acidic conditions within such stream reaches are not suitable for many aquatic species (Baker and Christensen 1991). With decreasing elevation, rates of Ca²⁺ supply by weathering increase and the effects of atmospheric deposition of strong acids on aquatic ecosystems and forest vegetation are diminished. These patterns are representative of montane landscapes throughout the Northeast (Driscoll 1991).

In summary, there are marked spatial variations in the Ca²⁺ dynamics of forest watershed ecosystems. Understanding variations in biogeochemistry across the landscape is critical to assessments of the response of watershed ecosystems to disturbance. Watershed-scale studies can take advantage of this characteristic by considering elevational variation in experimental design and measurements to help elucidate ecosystem processes.

A significant limitation of previous Ca²⁺ manipulation studies (i.e. CaCO₃ addition) has been uncertainty in the actual fate of the added material. A unique aspect of this proposed Ca²⁺ manipulation will be the use of Sr isotope ratios as a tracer to assess the transport, bioavailability and fate of added Ca²⁺. Strontium isotope ratios have long been used in dating geologic materials. More recently, they have been applied to the study of biogeochemical processes in natural ecosystems (Gosz and Moore 1989; Miller et al. 1993). Strontium isotopes hold great promise as a tracer because they are not fractionated by biological processes (Graustein 1989). Thus, variations in Sr isotopic ratios reflect variations in contributions from

Figure 2. Ecosystem pools (boxes) and fluxes (arrows) for Ca²⁺ for w6 at the HBEF. Average values in mol/ha or mol/ha-yr for periods indicated.. Values for 1964-69 are shown in A, while B shows the Ca²⁺ cycle for 1987-92. Net uptake values are based on the difference in biomass storage. (after Likens et al. 1997)

different sources. In forest ecosystems, the ultimate sources of Sr in are chemical weathering of minerals (including bedrock, till and soil) and atmospheric deposition. Strontium acts as an analog to Ca²⁺ because both are alkaline earth elements with similar ionic radii and the same valence (Elias et al. 1982; Jacks et al. 1989). At Cone Pond Watershed (10 km from the HBEF), Bailey et al. (1996) confirmed that ecosystem processes exhibited conservative Ca/Sr ratios, with the exception of some discrimination between the two elements in tree foliage.

Strontium and rubidium (Rb) are minor elements in the Earth's crust, with concentrations on the order of 10-1000 ppm (Turekian and Kulp 1956). ⁸⁷Rb decays to ⁸⁷Sr with a half-life of 5x10¹⁰ yr. Since ⁸⁷Sr is not radioactive, its concentration increases over time relative to ⁸⁶Sr which is non-radiogenic. The advantage to studying Sr is that its isotopic composition is variable, depending on the Rb/Sr ratio and age of its geologic source. Thus in small watersheds where the ⁸⁷Sr/⁸⁶Sr ratios of atmospheric inputs and weathering products are distinct and relatively constant over time, this method can be used to trace the transport of Sr through the ecosystem (Graustein 1989).

In this proposed experiment, we have a unique opportunity to monitor the dissolution of wollastonite and the fate of weathering products in the ecosystem via Sr isotopes and element ratios. The Sr isotope ratio of commercially available wollastonite (⁸⁷Sr/⁸⁶Sr = 0.70600; Blum, unpublished data) is distinct from that of rainwater (⁸⁷Sr/⁸⁶Sr 0.710) and bedrock (⁸⁷Sr/⁸⁶Sr 0.760) at Cone Pond Watershed (Bailey et al., 1996). Preliminary analyses of streamwater and groundwater from Hubbard Brook show that Hubbard Brook waters lie in the same range as those reported from Cone Pond (Bullen and Bailey, unpublished data). Note that differences in ⁸⁷Sr/⁸⁶Sr ratios in the 5th decimal place are routinely detectable by thermal ionization mass spectrometry.

The Sr isotope analyses will also yield insight into several important biogeochemical questions. For example, analyses of ⁸⁷Sr/⁸⁶Sr ratios in plant tissues can be used to estimate the proportion of the Ca²⁺ assimilated by plants which were derived from the added wollastonite compared to natural sources. In a similar manner, ⁸⁷Sr/⁸⁶Sr ratios in soil extracts can be used to study the relative importance of natural inputs and treatment inputs fo Sr (and Ca) to the soil exchange complex.

Previous studies of responses to additions of basic cations (mostly CaCO₃) and comparisons of naturally base-rich and base-poor sites suggest that base saturation is a fundamental controller of microbial biomass and activity in forest soils. Acid-base chemistry exerts direct control on microbial physiology and community structure, and perhaps more importantly, has an indirect influence on microbial biomass and activity through control of plant litter quality, faunal activity and soil organic matter quality.

The direct effects of additions of basic cations on microbial biomass and activity are to increase the specific activity of microbial cells (Ivarson 1977; Adams et al. 1978; Lohm et al. 1984; Zelles et al. 1978; Illmer and Schinner 1991) and to alter microbial community composition (Adams et al. 1978; Fritze 1991; Nodar et al. 1992; Frostegard et al. 1993). These changes, in combination with changes in the chemistry of recalcitrant soil organic matter, combine to increase the ability of microbes to utilize soil C pools (Persson et al. 1989). Many studies have observed increases in soil respiration in response to CaCO₃ addition (citations above). In the longer term, improvements in plant litter quality caused by additions of basic cations cause further fundamental changes in the nature and extent of microbial biomass and activity (Lukumbuzya et al. 1994; Fyles et al. 1994; Ouimet et al. 1996).

While additions of basic cations clearly increase microbial C processing in soil, effects on N dynamics are more complex and difficult to predict. Increased processing of soil organic matter can increase gross rates of N mineralization and immobilization (Hart et al. 1994), but may reduce net N mineralization if immobilization is stimulated more than mineralization (Nommik 1978; Nyborg and Hoyt 1978; Persson et al. 1989; Simmons et al. 1996a). Thus, the net effect is likely to be complex and variable in the long term. Effects on N dynamics are further complicated by effects of basic cations on soil fauna (described below), which can exert a strong influence on N mineralization dynamics (Ingham et al. 1985; Blair et al. 1994; Ronn et al. 1996).

In sum, the literature on additions of basic cations and comparison of base-rich and base-poor soils shows that there is a clear need for detailed, long-term studies of the consequences of supply of basic cations on microbial biomass and activity. A program of long-term measurements of litter quality, changes in soil organic matter C and N pools, gross and net rates of N turnover and microbial community structure are needed to understand how acid-base chemistry acts as a fundamental controller of microbial biomass and activity in forest ecosystems, and how these ecosystems will ultimately change in response to additions of basic cations.

Understanding the response of soil invertebrates to changes in acid-base chemistry is important because invertebrates are fundamental to decomposition and mineralization processes in soil (Seastedt 1984; Moore et al. 1988). Soil microarthropods influence decomposition rates and nutrient dynamics of decomposing litter (Witkamp and Crossley 1966; Seastedt and Crossley 1980, 1983; Blair et al. 1992) and perturbation of their communities may change decomposition rates and nutrient availability (Hagvar 1988; Heneghan and Bolger 1996). Furthermore, faunal influences on soil nutrient availability can positively affect plant growth (Setala and Huhta 1991; Setala 1995). Predictability and rapidness in the response of microarthropods to acidification makes them good biological indicators of forest acidification and decline (van Straalen et al. 1988; Hogervorst et al. 1993).

Because of the daunting complexity of soil invertebrate communities (Petersen and Luxton 1982), few studies have examined the response of soil invertebrates to changes in acid-base chemistry (Kuperman and Edwards 1997). Those investigations have focussed primarily on microarthropods (see below) and Enchytraeidae (potworms; Hagvar and Abrahamsen 1980; Abrahamsen 1983.). Certain microarthropod species are "calciophilic", whereas others are "acidophilic" (Hagvar 1990). For example, oribatid mites prefer more acidic soils, responding positively to acidification and negatively to base addition (Hagvar and Abrahamsen 1980; Hagvar and Amundsen 1981; Koskenniemi and Huhta 1986). On the other hand, most Collembolan species increased in abundance in response to base addition, although some species preferred more acidic soils (Abrahamsen et al. 1980; Baath et al. 1980; Hagvar 1984; Hagvar and Abrahamsen 1980; Kreutzer 1995).

Less information is available on the influence of acidification on macroinvertebrates. Millipedes (Diplopoda) appear to be insensitive to acidification (Esher et al. 1992) whereas centipedes (Chilopoda) increase in number in response to base addition (Theenhaus and Schaefer 1995; Schauermann 1985). Total numbers of macroinvertebrates in forest soil have been shown to decline in response to experimental acidification (Craft and Webb 1984) and to a gradient of increasing acidic deposition (Kuperman 1996), although populations of some species increased in response to sulfuric acid addition (Esher et al. 1992). Few studies have explicitly addressed the role of Ca²⁺, although it might be expected that populations of invertebrates with high Ca²⁺ demands (e.g. snails, centipedes, earthworms) would be affected by Ca²⁺ availability. Depletion of Ca²⁺ due to acidification in European forests has limited the breeding success in insectivorous birds (Graveland et al. 1994) and the Ca²⁺ content of caterpillars, beetles and aphids was higher on Ca-rich soils than on Ca-poor soils (Graveland and van Gijzen (1994).

The mechanisms that result in differences in the composition, diversity and dynamics of vegetation between base-rich and base-poor ecosystems are difficult to decipher because of a complex suite of historical and environmental factors, as well as feedback interactions, that influence vegetation. Similarly, the effects of gradual soil acidification (resulting from forest harvest and acidic deposition) on vegetation have proven difficult to document. The patterns and mechanisms of plant responses to an experimental shift in the base status of Hubbard Brook soils can contribute to our general understanding of vegetation-environment interactions in temperate forest ecosystems.

According to current theory and observation, plant diversity in forests varies as a result of spatial gradients in site fertility, disturbance and probably climate (Grubb 1986; Tilman and Pacala 1993; Houston 1994). Although diversity variation along the site fertility gradient may exhibit either a modal or a monotonically increasing pattern (Gentry and Emmons 1987; Grubb 1986) observations in the eastern deciduous forest biome suggest that herb species richness is markedly higher in base-rich than base-poor sites (Siccama et al. 1970; Graves and Monk 1985). Air pollution and soil acidification have been implicated in compositional shifts in herbaceous vegetation (Heil and Diemont 1983), including the ground-layer vegetation under forests (Nygaard and Abrahamsen 1991, Rodenkirchen 1992, Thimonier et al. 1992, Falkengren-Grerup 1995). However, the role of subtle natural processes in driving changes in forest herb communities may be difficult to discount, even in mature forests (Brewer 1980).

A recent, detailed comparison of the current composition, biomass and chemistry of herbs on w6 at HBEF (Siccama, unpublished data) with measurements from 1966 (Siccama et al. 1970) revealed several striking changes: i) the abundance (biomass, frequency) of several common, "acid-loving" herbs (Maianthemum canadense, Clintonia borealis, Lycopodium lucidulum, Dryopteris spinulosa, Oxalis montana) increased markedly since the mid-1960s; ii) several moderately common species apparently disappeared entirely from w6 during this interval (Arisaema triphyllum, Cinna arundinacea, several species of Carex, two species of Viola); iii) the abundance of several other common species declined markedly (e.g., Aster acuminatus, Cornus canadensis, several Rubus spp.); and iv) significant declines in foliar Ca of herbs have occurred (Likens et al. 1997). However, the role of Ca²⁺ and effects of depletion of basic cations from soil in driving these changes remains somewhat uncertain.

We anticipate that the ecosystem response to the experimental addition of wollastonite to w1 may be first evident through stream drainage. Calcium concentration has been strongly associated with stream productivity, with invertebrate biomass and productivity increasing with increasing Ca²⁺ concentrations (Egglishaw 1968; Krueger and Waters 1983). Experimental acidification of streams resulted in increased invertebrate drift, lower insect biomass, and increased periphyton (Hall et al. 1980). Sutcliffe and Hildrew (1989) noted that acidic streams have less invertebrate biomass and taxa richness, likely due to a combination of ionic stress caused by low concentrations of basic cations and decreased food quality caused by inhibition of microbial activity. Algal assemblage structure is also strongly affected by pH (Planas 1996).

There have been a few experiments examining the effects of CaCO₃ addition for mitigating acidification caused by acidic deposition (Rundle et al. 1995; Cirmo and Driscoll 1996; Schreiber 1996). Much of this research has focused on the biogeochemical response of streams to base addition (e.g. Cirmo and Driscoll 1996; Menendez et al. 1996), and fishes (e.g. Simmons et al. 1996b; Schofield and Keleher 1996); few studies have examined macroinvertebrate or microbial responses. In a Welsh stream experimentally manipulated by base addition (pH increase from 5 to 6.5), invertebrate taxa richness and abundance increased slightly, though the response was small when compared to a circumneutral stream (Rundle et al. 1995). Clayton and Menendez (1996) found similar results; acid-sensitive taxa increased during the study period, but there was no increase in abundance. No studies have examined the response of invertebrate production or biomass. Moreover, none of the base addition studies have examined lower trophic levels to determine whether increased food supply is a potential mechanism for responses by macroinvertebrates to changes in acidity. Stream acidity has been shown to lower microbial production and litter decomposition (Palumbo et al. 1987; Osgood and Boylen 1992). Because acid-sensitive taxa recolonize the stream after treatment (Clayton and Menendez 1996), there likely is a direct effect of pH on invertebrates. Hall et al. (1980) observed immediate decreases in macroinvertebrates in response to stream acidification, suggesting that physiological factors were responsible. However, it is possible that changes in microbial assemblages can affect invertebrate growth and production (Sutcliffe and Hildrew 1989; van Frankenhuyzen et al. 1985).

IV. Experimental Plan

The Hubbard Brook Experimental Forest (HBEF) is located in the southern portion of the White Mountain National Forest in Central New Hampshire. The climate is predominantly continental, characterized by long, cold winters and short cool summers (average temperature for January is -9^oC and for July is 10^oC). Annual precipitation is approximately 140 cm. The major bedrock map unit of the HBEF is the Silurian Rangely Formation, consisting of quartz mica schist and quartzite, interbedded with sulfidic schist and calc-silicate granulite. The soils are predominately Spodosols, Typic Haplorthods, derived from glacial till. The mean depth of the forest floor is 6.9 cm and the mean depth of the mineral soil is 50 cm. The overstory vegetation is northern hardwoods (Acer saccharum, Fagus grandifolia, Betula allegheniesis). Coniferous species are abundant at higher elevations (Abies balsamea, Picea rubens). Logging operations, ending around 1915-1917, removed large portions of conifers and better quality hardwoods. The present second-growth forest is even aged and composed of about 80-90% hardwoods and 10-20% conifers. Biogeochemical studies have been ongoing at the site since the early 1960's (Likens and Bormann 1995).

The proposed chemical manipulation will be conducted in watershed 1 (w1). Watershed 1 is 11.8 ha, and ranges in elevation from 490 -750 m. Hydrologic monitoring of w1 was initiated in 1956, measurements of stream chemistry have been made since 1966. The reference watershed for this proposed manipulation will be w6 (Likens et al. 1994, 1997).

Mineral-bound Ca is made available in HBEF soils through weathering of Ca-silicate minerals, such as plagioclase and hornblende (Likens et al. 1997). The proposed experimental application is designed to restore the base saturation of soil by an addition which mimics this natural weathering process as closely as possible but at a faster rate. Of various Ca-sources considered during pilot studies (Ca-acetate, Ca- bicarbonate, anorthosite), wollastonite (CaSiO₃) had the greatest potential because of the: i) minimum disruption of the microbial community (as observed for Ca-acetate); ii) ease in application (liquid application is not feasable due to the volume of water required to overcome solubility constraints); and iii) ability to increase base saturation because its weathering rate is among the fastest of the silicate minerals, while not releasing Al during weathering. Commercially available wollastonite has the added benefit of a Sr isotope composition (⁸⁷Sr/⁸⁶Sr = 0.70600) which is unique when compared to the natural range in ⁸⁷Sr/⁸⁶Sr in the Hubbard Brook region (Bailey et al. 1996), making it ideally suited as a tracer for the added Ca²⁺.

In pilot laboratory studies, wollastonite mixed with mineral soil from Hubbard Brook reacted at rates similar to those reported in other laboratory studies (e.g., Brantley and Chen 1995). However, dissolution in the field will likely be much slower due to imperfect wetting of mineral surfaces, accumulation of products and possibly fouling of the surfaces by organic and metal oxide coatings. Driscoll et al. (1996) found that 50% of the pelletized CaCO₃ added to the Woods Lake watershed, New York, dissolved over a two year period. Normalized to mineral surface area, this is equivalent to a dissolution rate three orders of magnitude lower than laboratory rates.

The amount of Ca²⁺ required to return the % base saturation from current levels ( ~10%) to the estimated levels of 50 yr ago (~19%) is 850 kg Ca/ha, equivalent to 30.2 tons of wollastonite over the area of w1. Based on: i) the difference between field and laboratory dissolution rates of CaCO₃ observed at Woods Lake Watershed; ii) the differences in laboratory weathering rates of wollastonite and CaCO₃, and iii) the difference in the particle sizes used at Woods Lake compared to material to be used in w1 (mean particle diameter = 3 mm), we estimate that about seven years will be required for the dissolution of the wollastonite in the ecosystem. Of the Ca²⁺ released by wollastonite dissolution under in pilot laboratory studies, 70% was retained on the exchange complex, whereas 30% was lost via drainage water. To account for loss from the watershed and provide a margin of error, the proposed dose for phase 1 will be 50% greater than the minimum required, or 45 tons. After 10 years, a second application will be made to further increase base saturation to the ultimate target level of 25%.

The application will be made in the fall of 1999. We believe the most cost-effective method of application with minimum disturbance to the watershed would be by helicopter. We envision that Joe Brigham Inc. of Concord, NH will do the helicopter application. Mr. Brigham is a licenced helicopter operator with the USDA Forest Service, has done other helicopter applications at Hubbard Brook (herbicide application to w2), and is routinely contracted to do sample collection and fish stocking in the White Mountains. Mr. Brigham has a 1 ton bucket, so the 45 ton dose could be loaded and applied in 1 day. For the application, wollastonite will be pelletized (1.5 - 4 mm) with a water-soluble binder (Driscoll et al. 1996). Pelletized material will help ensure that wollastonite will penetrate the overstory to the forest floor. The binder allows the pellets to disintegrate to the original grain size in the presence of moisture after the application.

During the pretreatment year of the study, Ca²⁺ and Sr concentrations and Sr isotopic composition of all major fluxes and pools within the watershed will be determined in order to develop Ca²⁺ and Sr mass balances for the watershed. Based on isotopic composition, via the method of Bailey et al. (1996), the Sr mass balance will be partitioned into two portions, quantifying the fate of Sr derived from atmospheric deposition separately from Sr derived from weathering reactions. Materials to be analyzed include atmospheric deposition (weekly collections), soil water (event and monthly collections), streamwater (event and weekly collections), bulk soil by genetic horizon, soil mineral separates, and biomass (sugar maple, American beech and red spruce wood, and foliage).

During the post-treatment phase, samples of the same types will be routinely screened for basic cation content and Sr isotope composition by inductively coupled plasma mass spectrometery (ICP-MS). This method allows quick and inexpensive determination of a large number of samples but only allows limited precision in isotope ratio determination. Based on changes in Ca/Sr ratio and Sr isotope composition, selected samples will be targeted for high precision analysis via thermal ionization mass spectrometry. These analyses will be used to quantify wollastonite dissolution and trace the fate of its weathering products. In addition to watershed scale analyses, subplots instrumented with lysimeters, aligned along a hydrologic flowpath, will be treated with artificial wollastonite manufactured with ⁸⁴Sr spike. Experimental manipulation of ⁸⁴Sr/⁸⁶Sr ratio will greatly increase sensitivity and longevity of the tracer allowing estimation of pedologic and hydrologic processes not amenable to study at the watershed-scale (see below). The fate of this tagged wollastonite will be monitored in vegetation soil, and drainage water.

At the HBEF, we use an integrated approach to investigate the response of watershed ecosystems to disturbances, including: i) longitudial measurements to assess spatial variation across the landscape; ii) process-level studies at the small-plot scale; iii) the use of tracers to quantify the transport and fate of materials; and iv) overall integration of the watershed study by mass balances (Likens et al. 1994; Romanowicz et al. 1996). This approach provides insight on important watershed processes such as: i) hydrologic flowpaths; ii) the role of vegetation composition in element dynamics, and iii) locations of element retention/release within the watershed. Following the addition of wollastonite to w1, we will use this approach to assess the impact of the Ca²⁺ addition through a comprehensive suite of measurements. Our monitoring initiative will include measurement of the major hydrologic fluxes, soil properties, microbial activity, phytosociology and plant tissue chemistry, and aquatic biota. Sampling will be conducted on both the whole-watershed scale and in intensively monitored plots (50-m x 50-m). Data from w1 will be compared to results from identical measurements in a reference area which includes w6 and the area immediately to the west of w6 (Bear Brook Watershed; BBW).

Precipitation is collected weekly at a clearing at the base of W6. Extensive analyses have revealed that there are no significant patterns in precipitation amount (Federer et al. 1990) or chemistry (Likens et al. 1977) with elevation or aspect within the study area.

Throughfall and soil solutions will be collected in four intensively monitored plots (50-m x 50-m) located throughout the elevational range of w1. Four nests of three funnel-type throughfall collecters will be placed randomly within each of the monitoring plots (Lovett et al. 1996). Throughfall from the three collecters in each nest will be composited, for a total of four samples per plot per sampling period. Samples will be collected within 24 hr of each rain event greater than 0.5 cm during the growing season. This approach is identical to the ongoing studies of throughfall in the BBW west of W6. In addition to this routine monitoring of throughfall, an additional 52 throughfall collectors will be deployed along two longitudinal transects in w1 prior to the helicopter application and for the summer season following the treatment. Throughfall from these additional collectors will be acidified in the field and analyzed for Ca²⁺. This intensive throughfall study will enable us to conduct a mass balance of Ca²⁺ through the canopy in order to accurately quantify the input of wollastonite-Ca²⁺ to the forest associated with the treatment.

Soil solution will be collected from zero-tension lysimeters placed below the Oa and Bh horizons, and within the Bs2 horizon (i.e., at the approximate depths of 6, 11, and 40 cm, respectively; Driscoll et al. 1988). The lysimeters were installed in the fall 1996 in the four intensive monitoring plots within w1. By the time of the wollastonite addition (1999), they will be well-equilibrated. At each lysimeter site, either duplicate or triplicate lysimeters were installed in each horizon. Soil solutions will be collected monthly throughout the year, and duplicate/triplicate samples composited in the field (Driscoll et al. 1988). Similar collections of soil solution have been made at three sites in the BBW since 1984.

Streamwater is collected weekly a few meters above the w1 outlet for chemical analysis. These samples, along with flow measurements recorded continuously at a V-notch weir, are used to estimate solute fluxes leaving the watershed-ecosystem. In addition, stream samples have been collected monthly since 1991 at four sites along the stream channel within w1 to assess longitudinal changes in stream chemistry within the catchment, which can be profound (e.g., Lawrence et al. 1986). This approach will allow us to detect both whole-watershed responses to the treatment, and also to identify sub-catchments which play a disproportionately important role in biogeochemical processes. For comparison purposes, streamwater from the w6 outlet has been collected weekly since 1963 (Likens and Bormann 1995) and monthly at five sites within the catchment since 1982 (Lawrence et al. 1986). All solutions will be analyzed for major solutes, dissolved organic carbon, dissolved inorganic carbon, total N, total Al, monomeric Al fractions, fluoride, dissolved silica, total phosphorus, soluble reactive phosphorus, ANC and pH using the methods described in Driscoll and van Dreason (1993).

In the summer of 1997, soil pits will be excavated at 48 randomly selected sites in w1. Fifteen of these pits will be excavated using a 0.5-m² frame to estimate soil mass for pool-size calculations (e.g., Huntington et al. 1988). Samples from the principal pedogenic horizons (E, Bh, Bs1, Bs2, C) will be collected from the faces of the exposed pits. In the summer of 1996, 15-cm x 15-cm forest floor "blocks" (Federer et al. 1993) were collected at 80 randomly selected sites within W1, and separated into Oi+Oe and Oa layers.

In the first phase of this study, the impacts of Ca addition on soil properties are likely to be limited to the forest floor and upper mineral soil. Therefore, our monitoring effort in the period covered in this proposal, will be focused on these layers. In the summer of 1998, the year prior to the wollastonite addition, we will again collect 80 forest floor blocks. After removing the forest floor, we will collect and composite two 10-cm core samples from the upper mineral soil. Identical procedures will be employed in the collection of 100 forest floor blocks and mineral soil cores from w6 in the summer of 1997. Forest floor and mineral soil core sampling will be repeated annually on w1 following the treatment. Complete characterization of soil chemistry using soil pits will be performed 7-8 years after the wollastonite addition, and will not be funded through this project.

All soil samples will be analyzed for pH (in H₂O and CaCl₂); loss-on-ignition (LOI); total C and N; exchangeable acidity (1M KCl extraction); exchangeable cations (Al, Ca, Mg, K, Na; 1M NH₄Cl); effective CEC (sum of exchangeable acidity and exchangeable bases); and total Ca (HCl-HF digestion). Specific chemical methods are detailed elsewhere (Johnson et al. 1991a,b; Johnson 1995; Johnson et al. 1995a,b; Johnson et al. 1997).

In this proposal, we request funds for measurement of the key variables that are most likely to show a rapid response to Ca²⁺ addition, including microbial biomass C and N content, soil respiration, potential net N mineralization, and nitrification. In each of the four intensively monitored plots, five replicate samples each of forest floor and the upper 10 cm of mineral soil will be collected three times per year, in early May, mid-July and early October. We expect to see marked increases in all these variables within five years.

Microbial biomass N and C content will be quantified as the inorganic N and CO₂ released in a 10 day aerobic incubation (25 ^oC) of a fumigated and re-inoculated sample (i.e., the chloroform fumigation-incubation method: Jenkinson and Powlson 1976; Voroney and Paul 1984). Soil respiration, potential net N mineralization, and nitrification will be quantified as the CO₂, total inorganic N, and NO₃^- produced in a 10-day aerobic incubation at 25 ^oC. Inorganic N will be measured colorometrically with a Perstorp flow injection analyzer. Carbon dioxide will be measured by thermal conductivity gas chromatography.

The sampling design for microbial measurements is identical to the ongoing measurement of a suite of microbial variables at four sites in the BBW reference area since 1994 (http://www.yale.edu/edex/groffman.html). Together the treatment and reference collections will result in a load of 240 samples per year.

Assessment of soil invertebrate communities will focus on micro- and macroarthropods. Both groups will be sampled three times per year (May, July, October) at the four intensively monitored plots in w1 and the BBW reference area. Paired plots in w1 will be selected according to similarity in overstory vegetation, slope, and aspect. The sampling locations will be adjacent to those used for microbial biomass and activity. For the collection of microarthropods, four soil cores (5 cm dia. x 10 cm) will be taken at each site (i.e., 16 cores each in w1 and BBW per date). Microarthropods will be extracted on high-gradient Tullgren-type extractors. The extracted animals will be counted and separated into the following major groups: i) suborders of Acari (mites) - Prostigmata, Mesostigmata (Gamasina, Uropodina), Oribatei, and Astigmata; ii) orders of other Arachnida; iii) suborders of Myriapoda, families of Collembola; and iv) orders or families of other Insecta. Population densities will be expressed as number per square meter to a depth of 10 cm. For interpretation of the results, soil from the cores will also be analyzed for gravimetric water content, pH (H₂O and CaCl₂), LOI, and exchangeable and total Ca²⁺.

Soil macroarthropods will be sampled using methodology similar to that described by Paquin and Coderre (1996). Four forest floor samples with an area of 12.5 x 25 cm will be collected in a block using a sampling frame. The sample blocks will be stored in bags until extraction of macroarthropods by handsorting. Extracted animals will be identified to order or family.

After separation, subsamples of major groups of both macro- and microarthropods will be oven dried, ground, ashed at 500 ^oC, and digested in hot, concentrated HNO₃. The resulting solution will be analyzed for Ca²⁺ using atomic absorption spectrophotometry.

The diversity, phytosociology, biomass, and chemistry of understory vegetation will be quantified in w1 using a combination of permanent plots and biomass harvest plots. Cover and frequency of herbs, shrubs and tree seedlings will be measured annually on a set of 200 1-m x 1-m permanent plots, in each 25-m x 25-m grid cell within w1. For the estimation of biomass and nutrient pools, we will also destructively sample understory vegetation in 200 1-m x 1-m plots following the same methods as used on w6 in previous studies (Siccama et al. 1970). Biomass is clipped aboveground, sorted by species, dried, weighed, ground, and stored for analysis. Both destructive and non-destructive sampling will be peformed in w1 and w6 in 1998 (pre-treatment) and 2001 (3-yr post treatment). Together, these approaches (permanent plots and harvest plots) will allow us to detect conclusively even subtle understory changes associated with treatment. Finally, during these measurement periods careful searches of w1 for rarer species, including mosses (Cleavitt and Fahey 1996), also will be conducted.

A complete census (species and diameter) of all trees > 10-cm dbh and a sub-sampling of trees with 2-cm < dbh < 10-cm was conducted on w1 in summer 1996. Similar resurveys will be conducted at 5-yr intervals over the duration of this long-term study. These measurements parallel our sampling program in reference w6. Overstory foliage of the dominant trees (sugar maple, beech, yellow birch) will be collected in the four intensively monitored plots in late summer of each year for tissue nutrient and specific leaf weight determination. At the same sites a suite of 48 litter traps (i.e., 12 in each intesively monitored plot) will be positioned for litterfall collection. Both of these sampling programs will parallel our reference collections in BBW. Biomass and chemistry of litterfall by species will be measured on these samples, and together with tree surveys and foliar chemistry sampling, will allow us to detect any gradual changes in foliage biomass, aboveground NPP, and nutrient and carbon retranslocation.

Biomass and depth distribution of fine roots will be measured in late spring (the point of annual minimum fine root biomass; Fahey and Hughes 1994) using soil coring methods. This sampling will be conducted in the intensively monitored plots in w1 and BBW. Sampling will be conducted prior to treatment (1998) and in post-treatment year 3 (2001). Soil cores are divided into forest floor and mineral horizons based upon our standard criteria and fine roots (< 1 mm) are sorted by hand (Fahey et al. 1988). Fine root chemistry will be measured annually in mid-summer for both forest floor and mineral soil roots of four dominant species in each of the elevation zones. Samples will be obtained by species, by sampling in monospecific groves (Fahey and Hughes 1994). Samples will be analyzed for Ca²⁺ and Al following cleaning and digestion procedures, and corrections for soil contamination as specified by Joslin and Wolfe (1989).

The response of stream biota to the experimental addition of wollastonite is a critical component of this ecosystem study. Funding for aquatic biota studies has been obtained through a grant to G.E. Likens from the A.W. Mellon Foundation. While the stream biota studies are briefly described here, no funding is requested for this work. We will examine stream biological response to watershed Ca²⁺ addition at several trophic levels. At the microbial level, we will measure how the Ca²⁺ addition affects bacterial production, fungal biomass, Ca²⁺ uptake, and algal taxonomic composition and biomass. We will also estimate secondary production of the entire invertebrate assemblage. Secondary production represents an informative method to examine invertebrates because it combines population-level changes with overall energy flux estimates. Because samples will be collected at bimonthly intervals, we will determine taxonomic changes in the assemblage through time and determine changes in invertebrate biodiversity.

Each autumn we will initiate an in-stream litter decomposition study to determine changes in microbial activity and associated decomposition. Bags of fresh sugar maple leaves will be placed in the w1 and w6 streams 50 m above the weir. We will use wide-mesh bags (1 cm) to allow invertebrate colonization. The bags will be sampled biweekly to calculate decay rate using a negative exponential model (Webster and Benfield 1986). Differences in leaf mass between the treatment and reference sites will be determined using analysis of covariance, using days in the stream as the covariate. We will examine fungal biomass in leaf packs using the method of Newell et al. (1988), whereby ergosterol is extracted from the leaf packs. Ergosterol is a fungal-specific sterol found in cell membranes of higher fungi in relatively constant amounts relative to living biomass. We will use the rate of accrual of fungal biomass to estimate net growth rates in leaves. Bacterial production will be determined using incorporation of leucine into bacterial cells (Kirchman 1993), modified for leaf disks (Suberkropp and Weyers 1996). We will also measure total Ca²⁺, N, and P in leaves from selected leaf bags to determine changes in nutrient concentrations.

Algal biovolume and taxonomic composition will be sampled bimonthly by brushing epilithon from rocks in both streams. Algae will be identified to genus, with diatoms to species to determine changes in taxonomic composition.

We will sample stream benthos in both streams approximately bimonthly throughout the year to estimate invertebrate secondary production. Six Hess samples will be taken from random spots in both streams. Samples will be poured onto 1 mm and 250 mm pore plates, sorted, identified and measured to the nearest mm. Insects will be identified to genus or, if possible, species. Biomass will be determined using length-weight regressions. Secondary production will be calculated for each taxon using the size frequency method corrected for cohort production interval (Benke 1984). Chironomid production will be calculated using the assemblage-level method of Huryn and Wallace (1986).

V. Project Management and Integration

While this proposed study is ambitious, we feel it is necessary to take an ecosystem approach to investigate the effects of soil Ca²⁺ depletion. We have considerable experience in large multidisciplinary, multiinvestigator studies. The wollastonite manipulation will be coordinated by Drs. C.T. Driscoll, S. Bailey and G.E. Likens and Mr. D. Buso. Measurements of bulk precipitation, throughfall, soil water, soil and streamwater chemistry response to the addition of wollastonite will be made by Drs C.T. Driscoll, C. Johnson, G.E. Likens and Mr. D. Buso. Drs. T.J. Fahey, T. Siccama and C. Johnson will oversee measurements of the response of forest vegetation to the wollastonite treatment. Dr. P.M. Groffman will supervise measurements of microbial activity. Dr. P. Bohlen will conduct measurements of the response of soil invertebrates to the experimental treatment. Dr G.E. Likens will supervise studies of stream biota. The Sr isotope component of this work will be conducted by Drs. S. Bailey , J. Blum, T. Bullen and C.T. Driscoll. Mr. D. Buso will be the project data manager. Data will be made available to the scientific community through the HBES data base within five years after data collection. If this proposal is funded, we will encourage the submission of "satellite" proposals by researchers interested in studying aspects of the ecosystem response to Ca²⁺ addition in more detail. Project coordination will be accomplished through the annual cooperators meeting of the HBES, quarterly project workshops and communication by telephone and electronic mail. A project web site will be established to provide an overview of the project, an update of project activities and facilitate data exchange.

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