M. PÉREZ–SUÁREZ
Instituto Potosino de Investigación Científica y Tecnológica, IPICYT
Camino a la presa San José 2055, San Luis Potosí, S.L.P.
Corresponding author; e–mail: marlin@ipicyt.edu.mx

M. E. FENN
United States Department of Agriculture Forest Service
Pacific Southwest Research Station, Forest Fire Laboratory, ]]> 4955 Canyon Crest Dr., Riverside, CA 92507. USA

V. M. CETINA–ALCALÁ, A. ALDRETE
Instituto de Recursos Naturales, Colegio de Postgraduados,
56230 Montecillo, México

Received March 5, 2007; accepted August 15, 2007

RESUMEN

ABSTRACT

Throughfall and soil chemistry were compared in two sites with differing atmospheric deposition: Desierto de los Leones National Park (high atmospheric deposition) and Zoquiapan National Park (low atmospheric deposition). Throughfall fluxes of NO₃–, SO₄^2–, Ca, Mg and K were compared under two canopy cover types: Abies religiosa Schl. (fir) and Pinus hartwegii Lindl. (pine), in comparison with sites without cover canopy, e.g. forests clearings. Throughfall fluxes decreased in the following order: fir > pine > forest clearing. Nitrogen balance under canopy of fir and pine resulted in negative values for net throughfall of NH₄⁺ at Desierto de los Leones and Zoquiapan, while NO₃–, only resulted in negative values under canopy cover at the low deposition site. With few exceptions, concentrations of total C, N and S, soluble SO₄^2–, and Ca²⁺ were higher in soil under fir canopies than under pine or in forest clearings. In polluted sites, the densely foliated fir canopies generally resulted in higher throughfall fluxes and soil accumulation of N, S and Mg compared to pine canopies or open areas. The elevated atmospheric depositions affect the functional process of forest ecosystem, particularly the throughfall and nutrients intern cycle, and these effects depend of the cover and present tree species.

Keywords: Throughfall, canopy leaf area, soil enrichment, México City air basin, atmospheric deposition.

1. Introduction

Forests downwind of México City are exposed to nitrogen (N) and sulfur (S) deposit (Fenn et al., 2002c) as well as high levels of ozone (Miller et al., 2002). Heavy metal concentrations in soil are also higher than upwind forest sites (Fenn et al., 2002b; Watmough and Hutchinson, 1999). Considerable evidence indicates that Pinus hartwegii Lindl. and Abies religiosa Schl. (sacred fir), the dominant coniferous species in the Desierto de los Leones National Park, are highly affected by air pollution (Alvarado et al., 1993; Rodríguez–Franco, 2002). P. hartwegii exhibits the classic symptoms of ozone injury, including chlorotic mottle, premature needle senescence and abscission, and canopy dieback. Sacred fir stands, in areas of the park called cemeteries, have suffered severe decline and mortality, beginning in the late 1970s (Alvarado–Rosales and Hernández–Tejeda, 2002). Experimental and circumstantial evidence suggests a role of air pollution in the decline of sacred fir. Manipulative field studies indicate that gaseous air pollutants, presumably photochemical oxidants such as ozone, cause visual foliar injury in sacred fir (Alvarado et al., 1993; Alvarado–Rosales and Hernández–Tejeda, 2002). Some have suggested that atmospheric deposition contributes to nutritional disorders of sacred fir (López–López et al., 1998), which may have a causal role in the forest decline.

Little data on deposition inputs of N and S in these forests have been published. Throughfall deposition of N and S was 18.5 and 20.4 kg ha^–1 yr^–1, respectively in an open stand of P. hartwegii in the Desierto de los Leones National Park, located southwest and downwind of México City (Fenn et al., 1999). By comparison, N and S deposition were considerably less (5.5 and 8.8 kg ha^–1 yr^–1, respectively) in a pine forest at Zoquiapan, an experimental forest to the east and upwind of México City. Sulfur deposition at Zoquiapan was slightly elevated because of volcanic SO₂ emissions from the Popocatepetl volcano (28 km south of Zoquiapan) that became active during the throughfall sampling period (Fenn et al., 1999). Thus, the effect of air pollution depends of species of tree and site condition. Deposition inputs have not been measured in sacred fir stands in the México City air basin, but deposition is expected to be higher to fir canopies than pine because of the much higher leaf area and stand density of A. religiosa compared to P. hartwegii.

The primary objective of this study was to determine the effect of canopy cover on ionic fluxes in throughfall, and on elemental and ionic concentrations in soil at high and low deposition sites within the México City air basin. We hypothesized that deposition fluxes in throughfall and nutrient and pollution enrichment in soil will vary in the order of decreasing canopy leaf area as follows: fir > pine > forest clearings. We also hypothesized that the nutrient and pollution enrichment effect of fir canopies on throughfall solutions and soil would be greater in the site with higher atmospheric deposition.

2. Materials and methods

2.1 Study locations and site description

Two sites in the Basin of México with differing levels of N and S deposition were included in this study (Fig. 1; Table I). The Desierto de los Leones National Park site (DL) is a highly–polluted site located in the direction of the prevailing winds from México City (Fenn et al., 2002a). Zoquiapan (ZOQ) is relatively low pollution site located upwind of México City and further away from the major urban areas than DL (Fig. 1). Soils at the study sites are classified as Andisols (Marín et al., 2002). The climate of the study sites is temperate sub–humid with a summer rain season and winter dry season. The moist easterly trade wind current brings convective rains into the Basin of México from May to October (Jáuregui, 2002).

2.2 Soil sampling and analysis

Soil samples were collected for nutrient and trace metal analyses from two study sites, also were analyzed for heavy metals (data reported previously in Fenn et al., 2006b). The soils at the study sites do not have a significant organic layer (approximately 1.5 cm), and it was removed, so only the mineral soil was sampled. At each site, composite soil samples of approximately 1 kg were collected at a depth of 0–5 cm and 5–15 cm from under the canopies of each of five A. religiosa and five P. hartwegii trees, and three samples from three adjacent clearings. All soils were sieved to 2 mm and the replicate soil samples from each site were then composited. Total C, N and S were determined by combustion analysis (Carlo Erba Instruments, Milan, Italy; Model NA 1500, Series 2). Trace metal concentrations (Ba, Li, Sr, Si, Fe, Ca, Mg, K and Na) and heavy metals (Pb, Ni, Co, Cr, Cu, Zn, Mn and Al) in the soils were determined by the microwave digestion technique described by Fenn et al. (2002b, c) as modified from Millward and Kluckner (1989). A single composite soil sample from the 0–5 and the 5–15 cm depth from each site was microwave digested using concentrated HNO₃ and HF according to USEPA method 3052 (USEPA, 1995). Total metal concentrations in the digests were determined by atomic absorption spectrophotometry (Fenn et al., 2002a).

Soils were also analyzed for NO₃–, NH₄⁺, and SO₄^2– . The soluble SO₄^2– fraction was extracted with water followed by two sequential extractions of the adsorbed fraction with 0.02 M KH₂PO₄ as described by Autry et al. (1990). The two phosphate extractions were combined to determine total adsorbed sulfate. Sulfate concentrations in the extracts were measured by ion chromatography (Dionex model DX–600, Sunnyvale, CA). Extractable NO₃– was determined in both aqueous and 2N KCl extracts. Ammonium was also determined from 2N KCl soil extracts. Nitrate and ammonium extracted with KCl were analyzed colorimetrically with a Technicon TRAACS 800 Autoanalyzer (Tarrytown, NY) and NO₃– in aqueous extracts was analyzed by ion chromatography.

2. 3 Bulk deposition and throughfall sampling

Throughfall and bulk deposition (deposition wet and dry, total in open areas) were collected in the DL and ZOQ sites over a 12–week period from August 1 until October 26, 2003, by which time the rainy season had ended. At each site, twenty P. hartwegii trees and twenty A. religiosa trees were randomly selected for throughfall sampling. As much as possible, trees were chosen so that the canopies of selected trees did not overlap. Under each tree a throughfall collector was placed midway between the main bole and the edge of the canopy drip line. The throughfall collectors consisted of a polyethylene funnel, 14 cm in diameter that was connected to a 4–L collection bottle covered with a dark plastic bag. The height placement of the funnel was 2.1 m above ground level. Six collectors were also placed in an adjacent open area for bulk deposition sampling. On a weekly basis the samples were collected and the volume of the solutions from each collector was measured. Sub samples of composite solutions from four throughfall collectors were combined to create five replicate samples per weekly sampling. Sub samples from pairs of bulk deposition samples were combined to obtain three replicate samples per weekly sampling. The composite samples were filtered with 0.45 µm pore size nitrocellulose membrane filters and stored under refrigeration until shipment to the USDA Forest Service, Forest Fire Laboratory in Riverside, California for analysis. Throughfall and bulk deposition solutions were analyzed for NO₃–, NH₄⁺, and nutrient cations (Ca, Mg and K) by ion chromatography using a Dionex DX–600 instrument. Ammonium concentrations were measured with a Technicon TRAACS 800 Autoanalyzer.

Because leaf area index (LAI) is an important parameter affecting atmospheric deposition fluxes to forest canopies, we estimated projected (one–sided) LAI under the pine and fir trees used to sample throughfall at DL and ZOQ. LAI was estimated with an indirect optical method (Pierce and Running, 1988). Canopy transmittance was measured with an integrating radiometer (Ceptometer, model SF–80, Decagon Devices, Inc., Pullman, Washington). Radiation measurements were taken in circular "sweeps" consisting of 16 measurements at each point where a throughfall collector was placed. Because the radiometer consists of 80 sensors, this resulted in 1280 point measurements of photosynthetically active radiation per throughfall collector location. From these data LAI values were calculated for each sampling point (Pierce and Running, 1988).

2.4 Statistical analyses

Statistical analyses were performed using SigmaStat statistical software, version 2.03 for Windows 95 (Jandel Scientific Software, San Rafael, California). Comparisons of throughfall deposition fluxes, pH and volumes between canopy cover types (fir, pine and open) were analyzed with one–way ANOVA followed by Tukey's all pairwise multiple comparisons procedures (p < 0.1). When the data failed normality tests, the one–way ANOVA on ranks followed by Dunn's multiple comparisons test was utilized. Comparison of net throughfall fluxes (the enrichment of element flux under canopy, throughfall–precipitation deposition) between fir and pine canopies (Fig. 3) were analyzed with paired t–tests.

3. Results

3.1 Bulk deposition and throughfall

Deposition in throughfall was higher at DL than ZOQ for NO₃–, NH₄⁺, SO₄^2–, Ca, Mg and K. Net throughfall deposition was negative at ZOQ for NO₃ collected under pine and fir, and even more for NH₄⁺. Sulfate collected under pine at ZOQ was marginally negative. At DL, net throughfall fluxes of NH₄4⁺ were strongly negative under both pine and fir canopies (Fig. 3). Net throughfall deposition of NO₃ under fir and pine canopies were 5.56 and 1.11 kg ha"¹ at DL compared to –0.16 and –0.29 kg haNO^–13 at ZOQ. Similarly, net throughfall deposition of SO₄² under fir and pine canopies were 4.98 and 2.36 kg ha–¹ at DL compared to 0.35 and –0.05 kg ha–¹ at ZOQ (Fig. 3).

Average pH values for throughfall under pine (p = 0.07) and fir (p = 0.001) were more acidic at DL (pH 5.23 for fir) than at ZOQ (pH 5.78 for fir). At both sites the pH of throughfall collected under pine trees was more acidic than solutions collected under fir trees or in forest clearings (Table II).

3.2 Effects of canopy cover and soil depth on soil chemistry

At DL total C, N and S concentrations in 0–5 cm soil were greatest under fir canopies and lowest under pine (Table III). However, in 5–15 cm soil from DL total C, N and S concentrations were higher, and C:N was lower in open areas than under pine or fir canopies. Unlike DL, ZOQ had the highest concentrations of total C, N and S in 0–5 soils under pine canopies, although differences among cover types were small.

Soil C: N ratios were lowest at DL, and were generally the highest at ZOQ (Table III). Surface soil from ZOQ had the lowest total S concentrations. The highest S concentrations were in 0–5 cm soils from DL, especially soil collected under fir canopies (1.12 and 0.36 g kg^–1 at DL and ZOQ, respectively). Concentrations of S in 5–15 cm soils were lowest at ZOQ and highest at DL (Table III). At both soil depths total N concentrations were lowest at ZOQ and highest at DL.

Adsorbed and soluble SO₄^2– concentrations in soil from open areas and soluble SO₄^2– in pine soils were similar at two sites (Table IV). Soluble SO₄^2– in soils under fir canopies was lowest both depths and in the 5–15 cm soils from DL. In 5–15 cm fir soil adsorbed SO₄^2– concentrations at ZOQ (57.88 mg kg^–1) were 1.6–2.3 times higher than at DL.

At both sites concentrations of Ca, and soluble SO₄^2– were higher in soil collected under fir canopies than soil from pine or open areas, except in 5–15 cm soils at DL where concentrations of Fe and K were highest in open area soils (Tables IV and V). Magnesium concentrations in soil differed little among cover types (Table V).

Concentrations of NH₄⁺ in fir soils were higher at both depths at DL than at ZOQ (Table IV). Potassium concentrations were similar (0.03–0.07%) at ZOQ and DL. Magnesium was the nutrient that was most enriched in soil at DL compared to the low deposition site, ZOQ. Magnesium concentrations were two to six times higher in soil from both soil depths at DL compared to ZOQ. Sodium concentrations in soil were 0.02–0.03% at both sites.

4. Discussion

4.1 Atmospheric deposition

Greater leaf area increased precipitation interception rates and low throughfall. However, in forests with high fog, the canopy increased fog interception rates and this fog is condensed and drips to soil. Thus, it can explain why measured rainfall volume at forest clearings is lower than under pine canopy at DL (Table II).

Greater deposition under fir canopies is presumably a result of the greater surface area of the fir canopies compared to pine canopies for capturing air pollutants. At DL and ZOQ, values for LAI of fir were 2.0 and 5.6 times greater than for pine (Table II). Others have reported that throughfall deposition of N and S decreases in the order of fir > pine > oak (Houdijk and Roelofs, 1991; van Ek and Draaijers, 1994). However, at both DL and ZOQ, deposition of NH₄⁺ was greatest in open areas, indicating high levels of NH₄⁺ consumption within fir and pine canopies. As a result, net throughfall fluxes of NH₄⁺ were negative for both tree species at both sites. Nitrate was also consumed by the canopy as evidenced by negative net throughfall values at ZOQ. At DL atmospheric deposition fluxes of NO_X to the canopy was much greater than canopy consumption rates, resulting in a positive net throughfall flux of NO₃–. Tree canopies appear to have also consumed SO₄^2–, as evidenced by a slightly negative value for net throughfall SO₄² flux under pine at ZOQ. This suggests that at least under pine canopies, throughfall flux of SO₄^2– is an underestimate of total SO₄^2– deposition, even during the wet season. It is generally accepted that throughfall deposition of SO₄^2– is a reasonable estimate of total S deposition, with minimal canopy retention of atmospheric S (Butler and Likens, 1995; Kovacs and Horvath, 2004; Lindberg and Lovett, 1992). Exceptions to this generalization would be forests exposed to elevated SO₂ concentrations resulting in canopy uptake and assimilation of SO₂ (Gay and Murphy, 1989; Granat and Richter, 1995; Manninen and Huttunen, 2000), and possibly areas with prolonged dry periods such as occur in the Basin of México (Fenn et al., 1999).

Deposition in throughfall of Ca, Mg and K was greatest under fir, intermediate under pine and lowest in open areas at DL and ZOQ. This is as expected because of the larger foliar surface area of the fir canopies. The largest source of these ions was likely leaching losses from foliage, in addition to washoff of deposition from canopy surfaces and wetfall deposition. In the San Bernardino Mountains of southern California concentrations of NO₃–, SO₄^2–, Cl^–, Ca, Mg, Na and K in throughfall and bulk deposition were greatest under white fir (Abies concolor Gord. and Glend.), intermediate under Jeffrey pine (Pinus jeffreyi Grev. and Balf.), and lowest in open areas (Fenn and Bytnerowicz, 1997). Throughfall fluxes of Ca, Mg and K were greater at DL than at ZOQ. The source of the greater cation fluxes at DL may be a result of the much greater throughfall fluxes of NO₃ and SO₄^2– at DL, with the basic cations functioning as counter balancing ions in the throughfall solutions. Higher fluxes of N and S at DL are a result of high atmospheric pollutant concentrations and greater precipitation volumes at DL. Calcium and Mg inputs were also higher in bulk deposition at DL than at ZOQ, at least partially because precipitation volumes were greater at DL. It may be that Ca and Mg salts of NO₃– and SO₄^2– are important sources of atmospheric deposition of N and S at DL.

At DL and ZOQ fluxes of NH₄⁺ in bulk deposition were similar to those of SO₄^2– but greater than bulk deposition fluxes of NO₃–. This begs the question as to whether total deposition of NH₄⁺ to the forest canopy may be greater than deposition of NO₃– and possibly similar to total deposition of SO₄^2– . This can not be determined from the throughfall data, because it is not known how much NH₄⁺ is consumed by the canopy. Total deposition of NH₄⁺ to the fir canopies at DL and ZOQ was likely greater than to pine canopies, even though throughfall fluxes under pine and fir were equivalent. Similarly, ionic concentrations of NH₄⁺ in throughfall were not significantly different under white fir and Jeffrey pine trees in the San Bernardino Mountains (Fenn and Bytnerowicz, 1997). In the present study, deposition of NO₃ in throughfall under fir canopies was double that under pine at DL, further suggesting that NH₄⁺ deposition to fir canopies was also much greater than to pine. Even though a large amount of atmospheric NH₄⁺ was retained by the forest canopy, atmospheric NH₄⁺ consumed by the canopy would still contribute to greater soil N enrichment under fir compared to pine (or under fir and pine compared to open areas) because the NH₄⁺ adsorbed or taken up by the canopy will eventually enter the soil N pool as litter fall or as soil organic matter derived from root turnover.

Precipitation and throughfall are major pathways in nutrient recycling, and they are an important source of nutrient input to forested ecosystems, two important processes functional for the ecosystem. Throughfall may be important in plant nutrition and soil fertility (Chapin et al., 2002). These effects, combined with the altered nutrient cycling regimes and the rates of soil processes resulting from forest growth, can affect site fertility and drainage water quality; aside from the magnitude and nature of the impacts have been shown to differ depending on the cover and species of tree.

4.2 Soil chemistry

4.2.1  Canopy covers effects on soil chemistry

In the current study, both soil NH₄⁺ and C concentrations were generally higher in pine and fir 0–5 cm soils than in open areas (except at ZOQ), probably a result of greater litter input compared to open areas. Likewise, at DL C concentrations in 0–5 cm soil were higher under fir canopies than under pine. Thus nutrient enrichment under fir was particularly pronounced in 0–5 cm soil at DL, the polluted site (Fenn et al., 2002a). Concentrations of C, N, S, soluble SO₄^2–, NH₄⁺, Li, Ca and K in 0–5 cm soil at DL were greater in soil under fir canopies than soil in open areas or under pine canopies. However, nitrate concentrations were an exception, as NO₃– concentrations in 0–5 cm soil were about five times greater in open soils than soils under pine or fir at DL (Table IV). Concentrations of total S and N in 0–5 cm soil at DL were 16 to 49% percent higher in fir soil than soil from under pine or in open areas. These results support the hypothesis that soil enrichment of N, S, Ca and K (and possibly other elements not measured in throughfall) under fir trees is greater than under pine or in open areas as a result of enhanced atmospheric deposition to fir canopies. On the other hand, for some constituents (i.e., soluble SO₄^2– and Ca), fir soils are more nutrient enriched than soils from under pine or in forest clearings, irrespective of atmospheric pollution levels. However, even for these constituents, the higher concentrations in fir soils could still be a function of the higher LAI of the fir trees.

A contrasting pattern of canopy cover effect was observed in the 5–15 cm soil from DL. Nitrate concentrations were 4 and 13 times higher in open soils than fir or pine soils in both sites ZOQ and DL (Table IV), it could be due to forests clearing often induce a higher microbiologic activity and mineralization rates. Higher nutrient concentrations in 5–15 cm open soils at DL could be due to lower nutrient depletion in the 5–15 cm zone from tree roots in the open areas, but we have no data to test this hypothesis, nor it is clear why this phenomenon was only observed at DL, the site most polluted (Fenn et al., 2002a), and with the highest annual precipitation and lowest annual average temperature (Jáuregui, 2002).

Concentrations of C, N, S, NO₃–, Fe and K were similar in the pine and fir soils, but they were consistently lower than in soils from open areas. The major nutrients C, N and S were 23–29% higher in open soils than fir or pine soils. K was 67% higher in open soil compared to fir or pine soils.

4.2.2 Site comparisons

Of the two study sites, soils at ZOQ reflect the least influence of atmospheric deposition on soil chemistry. However, concentrations of adsorbed SO₄^2– in soil were an exception. The highest absolute values for adsorbed SO₄^2– in pine soil at both depths were at ZOQ, and adsorbed SO₄^2– in 5–15 cm fir soils at ZOQ were two to three times greater than at the other three sites. In contrast, total S concentrations in soil at ZOQ were lower than at DL. The contrasting results for total S and adsorbed SO₄^2– at ZOQ may be explained by intermittent emissions of SO₂ from the Popocatépetl volcano which reawakened in 1993 after 70 years of quiescence. Volcanic activity increased in late 1994 and in the following three years Popocatépetl was one of the largest SO₂ sources in the world (De la Cruz–Reyna and Siebe, 1997). Sulfur deposition in throughfall (8.8 kg S ha^–1 yr^–1) in an open pine forest at ZOQ from March 1996 to March 1997 was higher than expected based on the dominant wind patterns in the México City air basin (Fenn et al., 1999). Throughfall fluxes probably underestimate total S deposition at ZOQ considering that little precipitation occurs for six months during the fall and winter and because ZOQ is a relatively low precipitation site compared to sites located south or west of México City. In summary, it may be that total S concentrations in soil at ZOQ have not increased, while adsorbed SO₄^2– has increased, because elevated S deposition at ZOQ is of recent and intermittent occurrence. It would likely take years of chronic S deposition to cause a clear signal of increased total S accumulation in soil; thus at DL, the most polluted site, both adsorbed SO₄^2– and total S in soil are elevated, as a result of chronic S deposition from fossil fuel emissions (Bravo–Álvarez and Torres–Jardón, 2002).

As observed in previous studies, soil C:N ratios are inherently low in the N–rich forest soils in the Basin of México (Fenn et al., 2002c, 2006a). As a result, differences in C:N ratios between high and low deposition sites are not large and are not always lower at more polluted sites. In this study, the highest C:N values (16.7–18.5) were at ZOQ a low deposition site. However, total N concentrations in soil were highest at the most polluted site (DL) and lowest at ZOQ.

4.2.3 Deposition effects on soil chemistry in relation to forest health

Soil pH was generally higher in soil under fir than under pine or in open areas (Fenn et al., 2006b), possibly because of the high base cation leaching rates from fir canopies. There was also a trend of lower soil pH at DL, the two high deposition sites, compared to the low deposition site (ZOQ), but differences were relatively minor (Fenn et al., 2006b). Likewise, the pH of throughfall and bulk deposition solutions was lower at DL than at ZOQ. High N and S deposition, which is usually associated with soil acidification, did not cause a significant effect on soil pH, presumably because of equally high levels of base cation deposition. As a result, soil acidification or base cation depletion does not appear to be a major factor in fir decline. The soils are sufficiently buffered and the high throughfall deposition of cations, N and S provides a relatively constant replenishment of these nutrients. Dendrochemical studies found elevated Pb and Cd levels in sacred fir tree rings since the 1960s at DL, but dendrochemical evidence (trends in wood Mn and base cation concentrations) indicate that soils under sacred fir at DL have not acidified, nor have base cation supplies diminished or available aluminum in soil increased during the past several decades when the fir decline problem became severe (Watmough and Hutchinson, 1999). These trends in soil chemistry and dendrochemistry are in contrast with declining fir and spruce stands in Europe and North America where base cation leaching from surface soil, soil acidification and Al mobilization are reported (Bondietti et al., 1990; Fenn et al., 2006a; Shortle et al., 1995).

Heavy metal concentrations in the same soils used in this study did not suggest a clear enrichment of heavy metals in soil under fir canopies as was found for N and S. Concentrations of Pb, Cu, Mn and Zn (Table VI) were, in general, slightly higher in fir soil than pine or open 0–5 cm soil, but concentrations of these same metals in 5–15 cm soil were higher in open areas than in fir or pine soils (Fenn et al., 2006b). From this it is not clear what effect pine or fir canopies have on levels of heavy metals in soil of the polluted study sites. Lack of an obvious fir enrichment effect on heavy metals, particularly in the deeper mineral soil, may be partially due to the low mobility of these metals that are likely immobilized by high organic matter concentrations in the upper soil horizon.

Heavy metal concentrations in soil at DL do not appear to be at phytotoxic levels for most plant species, although sensitive species, microbes or biological processes could be affected by the accumulated metals (Fenn et al., 2002b). In a recent study, the growth of eucalyptus seedlings planted into pine soil from Cumbres del Ajusco (AJ) National park (highly affected by air pollution) was extremely retarded (Fenn et al., 2006b; Perea–Estrada, 2003), but fertilization with N and P alleviated the growth reduction. The percentage of roots that were ectomycorrhizal, endomycorrizal and colonized by putative symbiotic endophytic fungi with abundant external mycelium was significantly lower for seedlings growing in soil from AJ and DL compared to ZOQ (Fenn et al., 2006b; Perea–Estrada et al., 2005). Thus, heavy metals or some other toxic factor in soil from AJ and DL may cause detrimental effects on mycorrhizae and possibly other symbiotic fungal/root associations (Perea–Estrada et al., 2005). These findings were based on soils collected under P. hartwegii, but soils under A. religiosa in the Basin of México may have the same characteristics, and potentially to a greater degree if the root cause is proportional to levels of atmospheric deposition. Clearly, further studies are warranted on the role of heavy metals or other possible unknown toxic factors in soil on mycorrhizal root development, and on possible relationships with the development of sacred fir decline. We are unaware of studies on the possibility of impaired mycorrhizal formation as a factor in fir decline.

5. Conclusions

Acknowledgements

The authors wish to thank Susan Schilling for producing Figure 1 and Víctor Perea, Timothy Blubaugh, Diane Alexander and Dave Jones for sample processing and chemical analyses.

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