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S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 a v a i l a b l e a t w w w. s c i e n c e d i r e c t . c o m w w w. e l s e v i e r. c o m / l o c a t e / s c i t o t e n v Tree-ring elemental concentrations in oak do not necessarily passively record changes in bioavailability Andrew R. Bukata⁎, T. Kurtis Kyser Queen's Facility for Isotope Research (QFIR), Department of Geological Sciences and Geological Engineering, Queen's University, Kingston, ON, Canada K7N 3N6 AR TIC LE I N FO ABS TR ACT Article history: Elemental concentrations in tree-rings from red and white oak trees at six sites across Received 2 April 2007 Southern Ontario, Canada, were assessed to determine whether they passively record Received in revised form changes in geochemical cycling in the presence of environmental stress. Periods of stress 6 September 2007 were defined as sustained periods with elevated δ13C values in tree-rings relative to Accepted 7 September 2007 atmospheric CO2 during the same period. In some trees, nutrient concentrations (Ca, Mg, Mn) were erratic during historic periods of stress while chemically similar non-nutrients (Ba, Keywords: Sr) and the anthropogenic pollutant Pb were not. Tree-ring concentrations of Ca and Sr were Dendrogeochemistry related to bedrock type and leachable concentrations in the soil. In contrast, tree-ring Carbon isotopes concentrations of Mg were not related to bedrock type, although Mg concentration in the soil Tree-ring leachate was. Tree-ring Mn, Ba and Pb concentrations were not related to bedrock type or LA-ICP-MS soil concentrations, but were inversely related to soil pH. Erratic behavior of nutrient elements during historic periods of stress suggests that some nutrient concentrations in the environment were not always passively recorded by tree-rings. © 2007 Elsevier B.V. All rights reserved. 1. Introduction Proxy records of pollution are necessary when direct instrumental monitoring of the environment is not available. Longevity, fixed location and visible annual growth rings make trees attractive as potential sentinels of environmental change. Chemical variations observed across tree-rings have led to the suggestion they can potentially be used to generate chronologies of changes in biogeochemical cycles (Watmough, 1997; Padilla and Anderson, 2002; Witte et al., 2004). For tree-ring elemental concentrations to be effective proxies for elemental bioavailability, trees must be passive monitors and the physiological effects on uptake of elements should be predictable. There must be minimal elemental mobility across growth rings and the elemental concentrations in the tree-ring must be proportional to bioavailability (Momoshima et al., 1995; Watmough, 1997), which may not always be the case (Watmough and Hutchinson, 2002). Failure to meet these conditions means that tree-ring elemental concentrations are not passive monitors and developing chronologies of environmental change based on tree-rings becomes difficult. There has been considerable debate as to whether elemental concentrations of tree-rings passively record bioavailability in their environment or whether concentration trends are dominated by tree physiology (Chun and Hui-yi, 1992; Smith and Shortle, 1996; Garbe-Schönberg et al., 1997; Watmough, 1999; Watmough and Hutchinson, 2003). The effect of stress on tree-ring elemental concentrations has not been systematically studied, but could help assess whether tree-ring chemistry passively records environmental conditions. This study examines the relationship between stress and tree-ring elemental concentration of nutrients (Ca, Mg and Mn), nonnutrients that are chemically similar to Ca (Ba, Sr) and the anthropogenic pollutant Pb. Despite chemical similarity, these ⁎ Corresponding author. Department of Geology, University at Buffalo, Buffalo, NY 14260-3050, USA. Tel.: +1 716 645 6800x6100; fax: +1 716 645 3999. E-mail address: [email protected] (A.R. Bukata). 0048-9697/$ – see front matter © 2007 Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2007.09.005 276 SC IE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 elements have been shown to display some dissimilar behavior in tree-rings including differences in radial mobility and uptake (Cutter and Guyette, 1993). If an environmental stressor affects geochemical cycling, similar behavior should be observed in both nutrients and chemically similar nonnutrients. Dissimilar response would indicate that environmental stress is affecting physiological functioning of the tree, which in turn differentially affects uptake and distribution of elements in the tree. Conversely, a lack of stress-related variation in uptake would be consistent with trees passively recording changes in their geochemical environment. Lead was analyzed as an element with an approximately known long-term pollution history in North America, with an increasing input function during the twentieth century that has been recorded in tree-rings with some success (Watmough, 1997, 1999; Watmough et al., 1998). The carbon isotopic composition of cellulose in tree-rings is affected by both the δ13C value of atmospheric CO2 and physiological functioning of the tree (Farquhar et al., 1982; O'Leary et al., 1992; Lajtha and Marshall, 1994; Panek and Waring, 1997). Atmospheric CO2 concentration has been increasing while its δ13C value has been decreasing since 1750, due primarily to the combustion of fossil fuels (Freyer, 1979; Freyer and Belacy, 1983; Leavitt and Long, 1989; O'Leary et al., 1992; Lajtha and Marshall, 1994; February and Stock, 1999). Farquhar et al. (1982) modeled the δ13C value of plant tissue with respect to atmospheric CO2 for C3 plants and found the plant δ13C values were controlled by photosynthetic pathway, diffusion related fractionation, and the ratio of the partial pressure of CO2 inside the leaf to atmospheric CO2. When stressed, decreases in stomatal aperture can occur (Kozlowski and Pallardy, 1997). This decreases isotopic fractionation resulting in higher tree-ring δ13C values (Freyer, 1979; Martin and Sutherland, 1990; O'Leary et al., 1992; Lajtha and Marshall, 1994; Panek and Waring, 1997; Sakata and Suzuki, 2000; Savard et al., 2002). As a result, positive deviations in tree-ring δ13C values from the long-term trend of decreasing δ13C values of atmospheric CO2 were used indicate periods of stress. This study is divided into three parts: (1) an examination of changes in tree-ring elemental concentration during periods of stress, (2) an examination of the relationship among tree-ring concentrations, site characteristics and soil leachate concentrations, and (3) an assessment of the cohesiveness of sitenormalized long-term non-nutrient elemental trends in treerings from sites with multiple trees analyzed. Together these data are used to assess whether tree-ring elemental concentrations passively record changes in their geochemical environment. We characterize periods of stress in individual red oak (Quercus rubra) and white oak (Quercus alba) based on increases in δ13C values in cellulose of tree-rings from trees at sites proximal and distal to urban areas in Ontario, Canada. Treering elemental concentrations during these periods of stress are compared to long-term trends to assess the extent to which variations in tree-ring chemistry may be physiological responses to stress. Tree-ring elemental concentrations were also considered with respect to soil characteristics such as soil pH and leachable concentrations of each element. Comparing tree-ring elemental concentrations to soil characteristics examines the extent to which recent tree-ring chemistries reflect soil characteristics and potentially may be reconstructed using historic tree-ring chemical compositions. In addition, the normalized long-term trends in elemental concentrations from multiple red oaks at three sites across Southern Ontario were compared to examine whether the normalization would be informative in developing sitespecific and regional chronologies. 2. Materials and methods Red and white oak trees were sampled at six sites across Southern Ontario, Canada (Fig. 1). These sites were selected to cover a broad range of soil and bedrock types as well as varying distances from urban pollution sources such as Hamilton, Toronto and Sudbury (Fig. 1 and Table 1). Forests at each site are mixed coniferous and deciduous remnant second growth stands. The sites are all located within the natural range for red and white oak, and while not the dominant species at every stand, they are abundant. When possible, white oak was sampled preferentially to red oak because it has preferred characteristics for dendrochemical analysis (Cutter and Guyette, 1993). Red oak is less preferred due to higher heartwood permeability and moisture content (Cutter and Guyette, 1993). Between three and ten dominant, visibly healthy trees were sampled at each site for potential chemical analysis. From each tree, eight cores were taken, two from each of four positions located at 90° intervals around the tree at breast height. Cores were taken using a 5 mm diameter increment corer. The corer was washed with ethanol followed by deionized water between trees and rinsed with deionized water between cores from the same tree. From these sampled trees, one to three trees from each site were selected for chemical analysis after an examination of the extracted core. Additional trees were sampled at each site for cross-dating purposes. These cores were air-dried, mounted and sanded with increasing grit number sandpaper until individual growth rings were easily visible. Cores from the same site were cross-dated according to the principles of Stokes and Smiley (1968). Distinctive ring-width patterns were identified in the cross-dated core at the same site and used to cross-date trees among sites prior to dissection for chemical analysis. For isotope analysis, cores from each position sampled around the bole were grouped and dissected into annual growth rings. The visibly distinct earlywood and latewood were separated in each growth ring and a sub-sample of the latewood from every even year was taken for carbon isotope analysis. α-Cellulose was prepared using a modification of the technique described by Loader et al. (1997) based on Green (1963). The carbon isotopic composition (δ13C value) of the α-cellulose fraction was measured using an Elemental Analyzer (EA) coupled to an Isotope Ratio Mass Spectrometer (IRMS) operating in Continuous Flow (CF) mode. EA-CF-IRMS analyses were performed using either a Carlo Erba NCS 2500 EA coupled to a Finnigan MAT 252 IRMS or a Costech ECS 4010 coupled to a Finnigan MAT Delta Plus XP in the Queen's Facility for Isotope Research. The δ13C values are reported in units of permil (‰) (Faure, 1986) relative to V-PDB. Using these techniques, NIST-19 returned a δ13C value of 1.95‰ and NIST-21 gave a δ13C value of − 28.1 ± 0.2‰. Replicate analyses S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 277 Fig. 1 – Map of Canada with inset map of Southern Ontario showing site locations and main pollution point sources of (Hamilton, Toronto, Sudbury). Arrows in inset indicate prevailing wind direction. of a laboratory α-cellulose standard (δ13C = − 24.3 ± 0.3‰) and selected samples indicate an uncertainty of 0.3‰ (2σ). Periods of stress were characterized as sustained intervals of elevated (N1‰) tree-ring δ13C values corrected for long-term changes in the δ13C values of atmospheric CO2. The temporal trend towards lower tree ring δ13C values due to the trend in atmospheric CO2 was removed by subtracting the δ13C values of atmospheric CO2 from the tree-ring δ13C value. Changes in tree-ring δ13C values relative to coeval atmospheric CO2 were compared by calculating Δδ13C values using the method of Freyer (1979). A time interval common to all sites was selected and the Δδ13C value was calculated as the difference between the measured tree-ring and the average δ13C value during that time interval. Intervals where the running average of Δδ13C Table 1 – Summary of the physical characteristics of the sampling sites in Ontario Site Location Bedrock type Soil type North Bay 46.35N 79.40W Granitic Gneiss Orthic Humo-Ferric Podzol Lake Opinicon 44.58N 76.32W Granitic Gneiss Monteagle Sandy Loam Podzol Burlington 43.28N 79.80W Shale/Mudstone Grey Brown Luvisol Peterborough 44.37N 78.30W Carbonate Orthic Melanic Brunisol Kingston North 44.33N 79.60W Carbonate Humic Gleysol Kingston East 44.26N 76.40W Carbonate Grey Wooded Podzol Soil pH (range) 4.62 (4.22–5.15) n = 10 5.07 (4.78–5.92) n=7 6.43 (5.78–8.27) n = 10 7.36 (6.62–7.93) n=9 6.87 (6.48–7.26) n=4 5.94 (5.69–6.65) n=3 Species Red oak Red oak Red oak Red oak White oak White oak 278 SC IE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 values exceeded 1‰ were identified as periods of stress. The onset and termination of the period of stress were defined as the beginning and end of the interval with elevated Δδ13C values. A change greater than 1‰ in δ13C value can be real and significant in plants (O'Leary et al., 1992) and as a result was chosen as a threshold value for stress definition. Sustained intervals of stress were used rather than individual measurements when the threshold value was exceeded to examine whether prolonged stress intervals had an effect on tree-ring elemental concentration. Short term stresses may lead to annual variations and appear as high frequency variability in the elemental concentration data. The exact timing of the onset and termination of these identified periods of stress are not rigid and as a result render the defined periods of stress qualitative to semi-quantitative. Individual cores were analyzed for elemental concentrations by Laser-Ablation Inductively Coupled Plasma Mass Spectrometry (LA-ICP-MS) as described by Kyser et al. (2003). All analyses were done with a Finnigan MAT Element High Resolution ICP-MS and a New Wave Research 213 nm Nd–YAG laser. The laser was operated in line or raster mode with a beam width of 300 μm and frequency of 20 Hz to measure each growth ring. Counts of each element were normalized to 13C to account for variations in ablation and transport efficiency to the plasma. Triplicate LA-ICP-MS analyses were done on at least three samples per tree and the results were compared to assure sample homogeneity and analytical reproducibility. To quantify the intensities from the LA-ICP-MS analyses, several tree-rings from each LA-ICP-MS run (60 rings total) were acid digested and their elemental concentrations determined by solution ICP-MS. Samples for solution ICP-MS analysis were selected to cover the widest range of machine response by LA-ICP-MS for the greatest number of elements. In a class-100 clean room, 0.020 to 0.100 mg of wood, all of the latewood remaining in the ring after LA analysis, was sealed in an acid washed Savillex container with 1–2 mL of high purity concentrated nitric acid. The container was placed on a hot plate (80 °C) for 3 days, five drops of ultra pure hydrogen Fig. 2 – a) Non-stressed tree from North Bay (red oak); b) Stressed tree from Lake Opinicon (red oak). The top row of figures is the carbon isotope (Δδ13C values), middle row is nutrients, and bottom row is non-nutrients. Hatched lines indicate the heartwood– sapwood boundary in each tree. Stressed time interval is indicated as a filled box. S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 peroxide added, and placed back on the hot plate for an additional day. Once clear, the solutions were evaporated to dryness in a laminar flow fume-hood and re-dissolved in 2% HNO3 with a 1 ppb indium internal standard. Analysis of standards and duplicates indicate a relative error of b5% for most elements. The concentrations measured by solution ICPMS were used to generate a 13C-normalized ion count per unit concentration for each element analyzed for each day of LA analysis. Individual conversion figures and equations were developed as in Watmough et al. (1998) and were plotted for each element in each core for each day run. Using the individual conversion equations, replicate laser-determined concentrations were within 20% for all the elements measured. The elemental concentrations were highly specific to individual trees (Figs. 2–4). Therefore, it became clear that comparing trends was of more interest than absolute elemental concentrations. To evaluate common temporal trends in the elemental concentrations of tree-rings for several trees from a specific site, we developed a normalization technique wherein an index value between 0 and 1 was calculated for each tree-ring for the time period common among all trees sampled at that site. For example, if three trees from one site covered the time periods 1860–1998, 1920–2004 and 1910–2000, the common time period for all three trees is 1920–1998. For each tree, the highest concentration measured in the tree-rings from 1920– 1998 was assigned an index value of 1, the lowest a value of 0, and the rest were given values between 0 and 1 scaled to their relative concentration using the following equation: Index value for year xðIx Þ ¼ ðconcx conclowest Þ= conchighest conclowest ð1Þ The calculated index value (Ix) for each tree-ring within the common time period at the site will be between 0 and 1 and those outside the common time period may be outside that range. Normalizing each core to a dimensionless unit also has the advantage that if the entire core is measured in a single laser ablation run without changing tune parameters and results are 13C-normalized to account for variations in ablation efficiency, the data need not be converted to concentrations prior to calculating index values. At each site, 3 to 10 soil samples were taken from the top 5 cm of the A-horizon. The samples were air-dried in a fumehood, organic debris was manually removed and the soil was passed through a 2 mm sieve. Soil pH was measured in a 1:1 soil/water slurry (Table 1). Leachable element concentrations in the soils were determined using a modification of the method described in Holk et al. (2003), wherein 0.5 g of soil and 5 mL of 2% HNO3 spiked with indium were placed in an ultrasonic bath for 2 h, centrifuged and the supernatant analyzed by ICP-MS. Such a harsh treatment was chosen to extract the highest concentration of elements possibly available to the tree. This fraction would include most soluble minerals as well as elements bound to amorphous oxides and humic matter. The strength of the relationships between soil pH, leachate and tree-ring elemental concentration was examined by linear regression in Microsoft Excel. 3. 279 Results Periods of stress were identified in all but one red oak from the North Bay site (Figs. 2–4). The periods of stress and the average δ13C value during the common interval among sites (1918 to 1930) and within the Burlington site (1936 to 1998) are summarized in Table 2. 3.1. Comparison of tree-ring elemental concentrations in the non-stressed and most stressed trees A red oak from North Bay does not record any prolonged periods of stress in its tree-ring Δδ13C values (Fig. 2a-i). The most pronounced change in nutrient concentrations in this tree occurs at the heartwood–sapwood transition, with a smaller change in non-nutrient element concentrations (Fig. 2a-ii and iii). There are shifts in Pb concentration independent of stress and in 1988 there is a spike in all elemental concentrations. In the heartwood of the nonstressed tree from North Bay, temporal trends in nutrient contents are similar to those of non-nutrients (Fig. 2a-ii and iii). The Δδ13C values in red oak from the Lake Opinicon site record a period of stress from 1940 to 1976 (Fig. 2b-i). These time intervals coincide with abrupt increases in the concentrations of nutrients Ca, Mn and Mg (Fig. 2b-ii). A concomitant increase in non-nutrient element concentrations during this period of stress is not observed, rather the tree-rings have Ba, Sr, and Pb concentrations that steadily increase after 1920 (Fig. 2b-iii). That there is no abrupt change in the concentrations of non-nutrients suggests the increase in nutrient concentrations represent a physiological response, rather than a change in geochemical conditions in the soil. 3.2. Stress coincident elemental variations in tree-rings During the period of stress in a red oak from Burlington (1946 to 1960; Fig. 3a-i) there is a decrease in Ca and Mn concentration (Fig. 3a-ii) unmatched by systematic changes in non-nutrient concentrations (Fig. 3a-iii). A second red oak from Burlington has two periods of stress, 1934 to 1938 and 1946 to 1958 (Fig. 3b-i). During the first period of stress the concentration of Ca increases and Mn decreases (Fig. 3b-ii) while the concentrations of non-nutrients Ba and Sr both increased (Fig. 3b-iii). During the second period of stress, overlapping spikes in Ca, Ba, Sr and Pb concentrations occur (Fig. 3b-ii and iii). Therefore, during both periods of stress in this tree there are shifts in the nutrient and non-nutrient concentrations. Nutrient and non-nutrient trends in the white oak from the Kingston East site have similar features during the period of stress from 1932 to 1966 (Fig. 3c). During the period of stress both the Ca and Sr concentrations reach their lowest values and begin to increase after the period of stress ends (Fig. 3c-ii and iii). Also, there are concomitant increases in Mn and Pb concentrations during the period of stress (Fig. 3c-ii and iii). Like the second red oak from Burlington (Fig. 3b), the white oak from the Kingston East site is recording similar shifts in both nutrient and non-nutrient concentrations during periods of stress. During a non-stressed interval between 1878 and 1896 in the Kingston 280 SC IE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 Fig. 3 – Stressed trees with elemental concentration shifts during period of stress from a) Burlington (red oak Burl4); b) Burlington (red oak Burl2); and c) Kingston East (white oak). The top row the figures is carbon isotope (Δδ13C values), middle row is nutrients, and bottom row is non-nutrients. Hatched lines indicate the heartwood–sapwood boundary in each tree. Stressed time interval is indicated as a filled box. S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 Fig. 4 – Stressed trees without elemental concentration shifts during period of stress from a) Peterborough (red oak); b) Burlington (red oak Burl3); c) Kingston North (white oak). The top row the figures is carbon isotope (Δδ13C values), middle row is nutrients, and bottom row is non-nutrients. Hatched lines indicate the heartwood–sapwood boundary in each tree. Stressed time interval is indicated as a filled box. 281 282 SC IE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 Table 2 – Average δ13C values during normalization periods for Δδ13C values and periods of stress δ13C value (‰; V-PDB) Avg. 1918–1930 Tree Burlington 2 Burlington 3 Burlington 4 North Bay Lake Opinicon Peterborough Kingston North Kingston East − 22.9 − 24.1 − 24.7 δ13C value (‰; V-PDB) Avg. 1936–1998 Period of stress −23.0 1934–1938 1946–1958 1956–1968 1946–1960 None 1940–1976 −22.6 −23.3 − 23.7 − 25.2 1932–1946 1952–1976 − 25.1 1932–1966 East white oak there are elevated nutrient (Ca, Mn and Mg) and non-nutrient (Ba and Sr) concentrations (Fig. 3c-iii). 3.3. No stress coincident variations in tree-rings elemental concentration While all three trees in Fig. 4 have periods of stress, in each case the threshold Δδ13C value of 1‰ is not greatly exceeded (Fig. 4a-i, b-i and c-i; Table 2). In particular, the Δδ13C in both the third red oak from the Burlington site (Fig. 4b) and the white oak from the Kingston North site (Fig. 4c) both have prolonged periods between 1950 and 1980 where the Δδ13C value is elevated near, but not exceeding the threshold value. Any systematic shifts in elemental concentration related to stress are not as clear in the three trees in Fig. 4 as they are in the Lake Opinicon red oak (Fig. 2b). After the period of stress in the Peterborough red oak (Fig. 4a) there is a concomitant increase in the Ca and Sr concentrations from 1950 to 1980, although the shift in Sr concentration is much smaller (Fig. 4a-ii and iii). Indeed the long-term Ca and Sr trends are similar to one another in both the red oak from Burlingon (Fig. 4b-ii and ii) and the white oak from Kingston North (Fig. 4c-ii and iii). 3.4. Soil leaches Elemental concentrations in soil leachates potentially reflect the amount available for uptake by trees. The highly individualistic behavior of trees with respect to total elemental concentrations (Figs. 2–4) and uptake, and possible relationships between heartwood and sapwood concentration confound a rigorous statistical treatment of soil chemistry here. Instead a qualitative examination of common features was done. Soils developed on carbonate bedrocks at Peterborough, Kingston North and Kingston East have higher pH and leachate Ca and Sr concentrations than low pH soils developed on granites at North Bay and Lake Opinicon (Fig. 5). High Ca and Sr concentrations in the soils are also reflected in relatively high Ca and Sr concentrations in the tree-rings. Calcium has higher concentrations in sapwood relative to heartwood concentrations, with the difference being more pronounced in trees over soils with high pH on carbonate bedrock (Fig. 5a). In all but the tree from Kingston East (pH = 5.94), heartwood and sapwood Sr concentrations are equivalent (Fig. 5b). Magnesium and Mn are both nutrients and are higher in concentration in sapwood relative to heartwood at all sites. Concentrations of Mg in soil leaches, like those of Ca and Sr, are higher in soils over carbonate bedrock sites and with higher soil pH, but sapwood Mg concentrations are similar at all sites, as are heartwood Mg concentrations (Fig. 5c). In contrast, soil leaches have similar Mn concentrations at all sites whereas low soil pH sites generally have higher Mn concentrations in both their sapwood and heartwood (Fig. 5e). Heartwood and sapwood concentrations are similar for both non-nutrient elements Ba and Pb at all sites (Fig. 5d and f). With the exception of the low pH site at North Bay that has elevated Ba concentrations in the soils, leachate concentrations of Ba are similar across the soil pH range of the sites sampled (Fig. 5d). However, tree-ring Ba concentrations vary with soil pH, with the highest concentrations on acidic soils. A linear regression of Ba concentration as a function of soil pH for the five sites with soil pHb 7 has an r2 = 0.85 for heartwood, an r2 = 0.83 for sapwood, whereas soil leach concentrations have an r2 = 0.30. This suggests Ba concentrations in tree-rings may be related to soil pH and not soil leachate concentration. Concentrations of Pb in the tree-rings also increase with decreasing pH and do not vary directly with Pb concentrations in the soils (Fig. 5f). 3.5. Long-term trends in non-nutrients Normalized index values were calculated for non-nutrients (Ba, Sr, Pb) from three trees sampled at Lake Opinicon, North Bay and Burlington (Fig. 6). Two of the three trees at Lake Opinicon have Ba and Sr tree-ring concentrations that have increased since 1940 whereas the third tree records decreasing Ba and Sr concentrations with time (Fig. 6a). All three trees have similar long-term Pb concentration trends towards higher concentrations since 1940 (Fig. 6a). Two trees from the North Bay site have identical temporal Ba index values, with a trend of decreasing values expected for most cations (Momoshima and Bondietti, 1990) until 1980, when the values increase (Fig. 6b). Index values for Sr similarly decrease until 1980 when they too increase (Fig. 6b). Two of the three trees have similar long-term trends towards higher Pb concentrations since 1940 and the third is consistent with the others until 1988, when the concentrations abruptly decrease (Fig. 6b). Trees from Burlington display the least temporal agreement among samples, except for Ba values which follow the long-term cation trend (Momoshima and Bondietti, 1990) (Fig. 6c). The Sr values are highly variable with no common temporal trend and Pb shows no distinct increase in concentration (Fig. 6c) despite the proximity to multiple point sources of pollution of this site (Fig. 1). 4. Discussion Environmental factors including climate, disease, pestilence and pollution can stress trees and affect carbon isotope compositions (Freyer, 1979; Martin and Sutherland, 1990; O'Leary et al., 1992; Lajtha and Marshall, 1994; Panek and Waring, 1997; Sakata and S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 283 Fig. 5 – Elemental concentrations in tree-rings and soil leachate concentrations plotted against soil pH for a) Ca, b) Sr, c) Mg, d) Ba, e) Mn and f) Pb. Legend for all panels is in the bottom right panel. From left to right on each panel, the sites are North Bay, Lake Opinicon, Kingston East, Burlington, Kingston North and Peterborough. Suzuki, 2000; Savard et al., 2002). Significant positive deviation in carbon isotopic composition can be used to define a period of stress but cannot establish the cause of the stress. As a result, variation in elemental concentration during these periods of stress may be directly recording the effect of the stressor in the geochemical environment, or may be recording a change in elemental uptake as a result of change in physiology. During periods of stress in some trees, nutrient concentrations in tree-rings become erratic, differing from long-term temporal trends of non-nutrients (Figs. 2b and 3a). This was particularly evident in the tree from Lake Opinicon (Fig. 2b). In this tree, the periods of stress are coincident with concentration increases in nutrients that are not shared by non-nutrient elements (Fig. 2b). These observations are consistent with stress-induced physiological changes in nutrient uptake or allocation occurring within the tree. In some instances, such as a red oak from Burlington (Fig. 3b) and a white oak from Kingston East (Fig. 3c), the stress-coincident shifts in nutrient concentration are accompanied by shifts in non-nutrient concentration, consistent with the stressor affecting the geochemical environment. In the Kingston East tree (Fig. 3c), there are concomitant shifts in nutrients and non-nutrients during a non-stressed interval (1878 to 1896). In other trees, there are no systematic shifts in elemental concentration coincident with periods of stress (Fig. 4). This may be due to chronic stress or may indicate that the level of stress is not sufficient to affect the physiological functioning of the tree or geochemical environment. Therefore, unless it is established that a tree is non-stressed, tree-ring nutrient concentrations cannot be considered exclusively passive monitors of elemental bioavailability. Differences between nutrient and non-nutrient behavior are more pronounced in tree-rings than in soil leachate concentrations or soil pH differences. With the exception of Ca concentrations at the low pH granitic sites (Fig. 5a), 284 SC IE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 Fig. 6 – Normalized non-nutrient concentrations (Pb top, Ba middle and Sr bottom row) for three trees from a) Lake Opinicon; b) North Bay; and c) Burlington. S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 sapwood has higher nutrient concentrations than heartwood regardless of soil pH or leachate concentration (Fig. 5a, c and e). Calcium and Mg have higher soil leachate concentrations above carbonate bedrock (Fig. 5a and c), and this appears to be reflected in elevated Ca concentrations in tree-ring concentrations (Fig. 5a). In contrast, Mg concentrations in tree-rings seem unrelated to soil leachate concentration or soil pH, with heartwood Mg concentrations generally low and comparable among sites, and sapwood Mg concentrations are high and comparable among sites (Fig. 5c). Both Ca and Mg concentrations in sugar maple tree-rings were shown to be related to soil leachate concentration (Watmough, 2002). The lack of relationship between soil leachate concentration and tree-ring concentration of Mg suggests oak tree-rings do not passively record the amount of this element available in the soil for uptake; rather is controlled by the physiology of the tree. Tree-ring Mn concentrations follow a pattern opposite that of Mg. Soil leachate Mn concentration is similar among sites, but tree-ring Mn concentrations vary with soil pH (Fig. 5e). Higher tree-ring Mn concentrations occur in soils with low pH, suggesting that soil pH affects tree-ring Mn concentrations, as has been reported in previous studies (Watmough, 1997, 2002). However, attempting to develop chronologies of soil pH using tree-ring Mn concentration might be hampered by potentially high radial mobility of Mn (Cutter and Guyette, 1993) and a strong physiologically-controlled trend seen for nutrients in oak with a decrease to the heartwood–sapwood boundary followed by an increase to the outermost sapwood ring (Figs. 2–4). Concentrations of Pb in oak tree-rings seem more related to soil pH than soil leachate concentrations or bedrock type (Fig. 5f), similar to previous observations in sugar maple (Watmough, 2002). Soils with low pH have higher tree-ring Pb concentrations than do soils with higher pH, independent of leachate concentration. There is excellent agreement among temporal trends in Pb concentrations in the three trees analyzed at Lake Opinicon (Fig. 6a), a site distal to urban areas of Southern Ontario, but with soils having low pH. Increasing Pb concentrations in these trees begin in 1940 and continue until the mid1990s, and the good agreement among trees suggests it may be recording a site-wide process. Two of the three trees sampled at North Bay, also a site with soils of low pH (Table 1), have temporal Pb concentration trends similar to those at Lake Opinicon. The third tree at North Bay shows the same trend until 1988, when the concentration suddenly decreases (Fig. 6a and b). The cause of this intriguing decrease is not clear. In contrast, the high variability and lack of agreement among treering Pb concentration trends at the Burlington site (Fig. 6c) suggest that either the circum-neutral soil pH is making the site unsuitable for developing tree-ring based Pb pollution histories or competing processes at the site are causing a variable signal to be recorded. Proximity of the Burlington site to pollution sources likely resulted in the relatively high soil leachate Pb concentrations, but this is not reflected in tree-ring Pb concentrations (Fig. 6f). The low radial mobility of Pb (Cutter and Guyette, 1993) suggests tree-ring Pb concentrations can be used to reconstruct site-wide changes in geochemical processes, but concentration trends may be complicated by soil pH. Tree-ring Ba concentrations seem independent of bedrock type or soil leachate concentration, but are inversely related to soil pH (Fig. 5d), and thus may record an environmental signal. 285 For example, barium concentrations in trees measured at the Lake Opinicon site, with low soil pH, have increasing or constant trends since 1950 (Fig. 6a). The inverse correlation between tree-ring Ba concentrations and soil pH (Fig. 5d) suggests the temporal trend in these trees may be recording the well-documented increased acid deposition in Canada throughout the second half of the twentieth century (Environment Canada, 2004). Momoshima and Bondietti (1990) demonstrated that tree-rings behave like ion exchange columns with the number of binding sites per ring decreasing radially from the pith. This can result in a trend of decreasing tree-ring cation concentration from pith to outer most growth ring, as is seen in the tree-ring Ba concentrations of the Burlington trees (Fig. 6c). Tree-ring Ba concentrations at North Bay similarly decrease outwards from the pith until 1960 when concentrations increase (Fig. 6b). This may indicate a change in Ba availability, possibly related to increased acid deposition since 1960 (Fig. 6b). Given the low radial mobility of Ba (Cutter and Guyette, 1993), tree-ring Ba concentration may be used to reconstruct historic changes in soil pH or acidic deposition. Barium and Sr are both alkaline earth elements and their similar chemical behavior is likely responsible for the similarity of their temporal trends at both the Lake Opinicon and North Bay sites (Fig. 6a and b). However, the agreement among temporal Sr concentration trends is not as good as for Ba at both sites. Lower soil leachate Sr concentrations and a lack of relationship between soil pH and tree-ring Sr concentration (Fig. 6b) may have combined to make tree-ring Sr concentration shifts less sensitive to environmental changes than Ba at Lake Opinicon and North Bay. The lack of a common trend of Ba and Sr in trees from Burlington is consistent with no site-wide environmental change recorded in tree-ring Sr concentrations. The use of a site normalization procedure for tree-ring elemental concentration allows for the examination of potential site-wide responses to environmental change recorded by tree-rings. Changes in nutrient concentration can be examined in treerings, but without knowledge of the stress-status of the tree, any changes not matched by chemically similar non-nutrients may be due to physiological stress in the tree. The effects of stress on tree-ring elemental concentrations may be partially mitigated by analyzing multi-year segments of the tree rather than individual rings. The potential effect of stress on elemental concentration would be averaged over the segment and minimized. The potential effect of stress and site characteristics on tree-ring chemistry suggests that background or control trees should be selected from sites that are as similar to the study area as possible, differing in only the single variable being examined such as in Witte et al. (2004) and Sheppard et al. (2007). Replication will help establish the veracity of any temporal patterns observed. Our results suggest that normalizing concentrations to a dimensionless index value can be useful in detecting common patterns among trees even when absolute concentrations are quite different. 5. Conclusions While nutrient concentrations in tree-rings from oak may be used in pollution studies, they must be used with caution as concentrations can be affected by physiological stress unrelated 286 SC IE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 0 ( 2 00 8 ) 2 7 5–2 86 to the pollution. 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