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Detecting changes in insect herbivore communities along a pollution gradient
Environmental Pollution 143 (2006) 377e387 www.elsevier.com/locate/envpol
Detecting changes in insect herbivore communities along a pollution gradient Michele Eatough Jones*, Timothy D. Paine Department of Entomology, University of California Riverside, Riverside, CA 92521, USA Received 13 September 2005; received in revised form 10 December 2005; accepted 14 December 2005
Differences in insect herbivore communities were associated with an ambient air pollution gradient in the mixed conifer forest outside the Los Angeles area. Abstract The forests surrounding the urban areas of the Los Angeles basin are impacted by ozone and nitrogen pollutants arising from urban areas. We examined changes in the herbivore communities of three prominent plant species (ponderosa pine, California black oak and bracken fern) at six sites along an air pollution gradient. Insects were extracted from foliage samples collected in spring, as foliage reached full expansion. Community differences were evaluated using total herbivore abundance, richness, Shannon-Weiner diversity, and discriminant function analysis. Even without conspicuous changes in total numbers, diversity or richness of herbivores, herbivore groups showed patterns of change that followed the air pollution gradient that were apparent through discriminant function analysis. For bracken fern and oak, chewing insects were more dominant at high pollution sites. Oak herbivore communities showed the strongest effect. These changes in herbivore communities may affect nutrient cycling in forest systems. Ó 2005 Elsevier Ltd. All rights reserved. Keywords: Nitrogen deposition; Pinus ponderosa; Quercus kelloggii; Pteridium aquilinum; San Bernardino Mountains
1. Introduction Human activities are widely reported to impact ecosystems on both local and global scales. Long term impacts can include changes in biogeochemical cycles, primary productivity, altered biotic interactions and decreased ecosystem resilience (Chapin et al., 2000; Naeem and Li, 1997; Taylor et al., 1994). Loss of species is assumed to be one of the driving mechanisms for changes in ecosystem function (Chapin et al., 2000). Monitoring species richness or individual species abundance has been promoted as a key method for detecting environmental impacts in many conservation and management programs (USDA Forest Service, 2002).
* Corresponding author. Tel.: þ1 951 827 4488; fax: þ1 951 827 3086. E-mail addresses: [email protected] (M. Eatough Jones), timothy. [email protected] (T.D. Paine). 0269-7491/$ - see front matter Ó 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2005.12.013
Insect herbivore communities may be good indicators of ecosystem changes. Insect herbivore populations can respond to changes in plant nutrients (Mattson, 1980), plant defensive chemistry (Rosenthal and Berenbaum, 1992; Feeny, 1976), plant growth characteristics (Mopper and Simberloff, 1995), or plant community structure (Jefferies and Maron, 1997; Brown et al., 1988). Insect herbivores also play critical roles in ecosystem processes. They influence nutrient cycling (Reynolds and Hunter, 2001; Belovsky and Slade, 2000; Schowalter, 1995) and provide food for higher trophic levels. Changes in the population density or structure of the herbivorous insect community can affect plant productivity, competition between plants, and plant community composition (Ritchie et al., 1998; Jefferies and Maron, 1997; Janzen, 1987). Major anthropogenic impacts affecting forest systems include habitat fragmentation, invasive introduced species, global warming, increased atmospheric CO2, and air pollution.
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Although air pollution affects many ecosystem types worldwide, forests in particular may be vulnerable to air pollution effects (Fowler et al., 1999). Ozone and nitrogen deposition are considered to be the two most important pollutants affecting forest ecosystems in North America (Fowler et al., 1999; Taylor et al., 1994). Atmospheric nitrogen deposition can acidify forest soils and promote leaching of base cations which may cause nutrient imbalances in plants that contribute to forest decline (McNulty and Aber, 1993). Nitric acid deposited on foliar surfaces may directly damage the plant cuticle resulting in increased loss of water and metabolites and greater susceptibility to insects or pathogens (Bytnerowicz et al., 2001). Deposition of atmospheric nitrogen can also act as a fertilizer, stimulating forest growth and increasing plant nitrogen content (Vitousek et al., 1997). Elevated nitrogen in plant tissues has been associated with increased insect performance and greater damage to plants (Kyto et al., 1996; Waring and Cobb, 1992; Lightfoot and Whitford, 1990; Skeffington and Wilson, 1988; Bryant et al., 1983; Mattson, 1980). Alternatively, higher nitrogen in plant tissues may lead to deterrence of insect herbivores through mechanisms including increased nitrate in plant tissues (Jansson and Smilowitz, 1985; Manglitz et al., 1976) and increased resin content in pines (Kyto et al., 1996). Ozone damaged plants may have higher concentrations of free amino acids, soluble proteins and soluble sugars that increase plant nutrients available for insects (Tingey et al., 1976; Barnes, 1972). Most chewing insects studied showed increased feeding and faster growth rates on ozone fumigated foliage (Heliovaara and Vaisanen, 1993). However, results for aphid feeding and performance were more varied. Other factors may also mitigate herbivore responses to ozone-stressed plants. In long-term population surveys, Percy et al. (2002) found no change in aphid performance on ozone fumigated trees, but aphid abundance increased, presumably through the negative effects of ozone on natural enemies of aphids. 1.1. Air pollution in southern California forests Atmospheric pollutants, particularly nitrogenous compounds and ozone are transported inland from urban areas of the Los Angeles basin, along the western slopes of the San Bernardino Mountains by prevailing weather patterns (Miller, 1992). Concern with air pollution damage in the San Bernardino Forest began as early as the 1960s when ozone injury on pine needle tissue was observed (Cobb and Stark, 1970; Miller et al., 1963). The wildland ecosystems surrounding Los Angeles receive the highest levels of nitrogen deposition recorded in North America (Fenn et al., 2003a). Concentrations of tropospheric ozone are also high in this region (Grulke and Balduman, 1999). Chronic exposure to air pollutants has affected plant growth and other plant characteristics (Grulke, 1999; Grulke et al., 1998; Miller and Rechel, 1999; Bytnerowicz and Fenn, 1996). This may potentially affect performance of individual plant species, competitive relationships within the plant community, and the structure and abundance of organisms in other trophic levels (Fenn et al., 2003b).
A well-established horizontal pollution gradient has been described in the San Bernardino Mountains with decreasing ozone and nitrogenous pollutant concentrations from west to east (Miller and Rechel, 1999; Miller et al., 1986). Air pollution impacts to plants have been established along this gradient. Ponderosa pine (Pinus ponderosa Laws) trees from highly polluted western areas have shown greater symptoms associated with ozone exposure than trees from less impacted eastern areas, including decreased growth (Miller, 1992), reduced root mass and belowground allocation of carbohydrates (Grulke et al., 1998), and chlorotic mottling of needles and reduced needle retention (Miller, 1973). Ponderosa pine is among the most susceptible plant species to ozone exposure in the San Bernardino Mountains (Miller et al., 1983), and many have already died, most likely due to a combination of stresses, including oxidant injury, bark beetle attacks and disease (Pronos et al., 1999). Total foliar nitrogen and nitrate were higher in bracken fern (Pteridium aquilinum var. pubescens Underw.), California black oak (Quercus kelloggii Newb.) and ponderosa pine from a western site compared to an eastern site (Fenn et al., 1996). We examined arthropod herbivore communities of three prominent plant species along the ambient pollution gradient in the San Bernardino Mountains. Although it is not possible to rule out effects from other environmental factors in any ambient gradient study, the strong history of documented air pollution effects along this gradient make it an ideal location to examine insect community impacts. We hypothesized that high atmospheric inputs of ozone and nitrogen deposition would lead to alterations in the composition of the insect herbivore communities. To assess if there were changes in herbivore communities along the gradient we used the traditional measures of richness and diversity, and also a multivariate analysis to examine community structure. We evaluated the performance of these different methods for examining herbivore communities and which methods of community comparison gave the clearest and most reproducible results from year to year. 2. Methods 2.1. Study locations Study sites were selected from permanent plots established in the early 1970s by the USDA Forest Service Pacific Southwest Research Station for long-term air pollution monitoring (Arbaugh et al., 2003). Sites were chosen along the west-east air pollution gradient in the mixed conifer zone (1300 m to 2700 m elevation) of the San Bernardino Mountains. Sites selected for this study range from 1500 m to 2300 m (Table 1). The dominant overstory species are ponderosa pine, California black oak and incense cedar (Calocedrus decurrens (Torr.) Florin.) with a minor component of sugar pine (Pinus lambertiana Dougl.) and white fir (Abies concolor Gord. & Glend.). Jeffrey pine (Pinus jeffreyi Grev. & Balf.) becomes more common at higher elevations, while incense cedar becomes less common. Bracken fern forms a dense understory in many locations. Additional details on climate, soils and vegetation are reported in Miller and McBride (1999). Of our study species, ponderosa pine is the most sensitive to ozone exposure (Miller and Rechel, 1999), while bracken fern has shown the greatest increase in foliar nitrogen content along the gradient (Fenn et al., 1996). All sites were subjected to drought conditions during this study. Annual precipitation in the San Bernardino
M. Eatough Jones, T.D. Paine / Environmental Pollution 143 (2006) 377e387 Table 1 Summer air pollution concentrations and geographical characteristics for sites in the San Bernardino Mountains Relative Elev. Distance O3 a NO2a HNO3a pollution (m) from L.A. (ppb) (ppb) (mg/m3) exposure (km) BL GV
Bluff Lake Green Valley Creek BF Barton Flats DW Dogwood BP Breezy Point CP Camp Paivika
Low Low
2320 117 1950 106
59.3
1.4
2.6
Low High High High
1900 125 1725 94.5 1525 86.1 1600 84.7
55.9 59.4 63.2 62.8
3.1 7.7 5.4 6.9
2.0 3.7 5.3 4.8
Air pollution data was not collected at BL during the study period but exposure is historically low. a Average atmospheric concentration determined by passive monitoring (Alonso et al., 2002; Bytnerowicz et al., 2002), June 7 through August 14, 2001. (Atmospheric Deposition Group, USDA PSW Research Station).
Mountains was below half-normal for four years beginning in 1999 and continuing through the study period (California Department of Water Resources, public information). Three western sites had high pollution inputs: Camp Paivika (CP), Breezy Point (BP) and Dogwood (DW). More eastern sites at Barton Flats (BF), Green Valley Creek (GV) and Bluff Lake (BL) had lower atmospheric inputs (Fig. 1). Elevation and precipitation also vary along the gradient, but pines along the gradient showed a strong east to west trend for increased foliar injury indicated by forest pest management index scores (Miller and Rechel, 1999). Recent air pollution concentrations suggested two contrasts among sites (Table 1). Data on nitrogen oxides suggested categorizing BF, GV and BL as low pollution and CP, BP and DW as high pollution sites. However, ozone concentrations at DW were more similar to those recorded at the low pollution sites. Ozone data suggested a contrast between CP and BP as high exposure sites and DW, BF, GV and BL as low pollution sites. We evaluated both these contrasts in our analysis of herbivore community characteristics. Contrasts were designed using the study sites as the unit of replication, with herbivore
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communities analyzed separately each year to mitigate lack of independence of replicates from resampling in the same locations each year.
2.2. Sample collection Insect and foliage collections from oak, pine and fern were taken as foliage reached full expansion in 2001 and 2002. Collection times were late May into June for fern and oak, and July for pine. Foliage samples of three 30 cm long branch tips were collected at mid-crown height from 10 oak and 10 pine at each site. Pine foliage included current year’s needles and older annual whorls on the branch. Foliage was removed using a pole-pruner with a basket lined with a 125 l trash bag attached below the pruner. Foliage was collected in a bag so that the bag could be closed immediately after collection to minimize loss of insect fauna through handling the sample. The above ground growth of bracken fern was bagged and removed from ten 0.5 m 0.5 m areas randomly selected at each study site. Collected foliage was stored on ice for transport and stored in an ultracold (65 C) freezer in the laboratory. Insects were manually removed from the foliage, and identified to family and morphotaxa. Morphotaxa were assigned based on visually distinct appearances. Common morphotaxa were collected from foliage near plot areas and reared in the lab so that all stages could be identified and grouped. Herbivore abundance was calculated as insects/gram fresh weight foliage for pine and fern and as insects/leaf for oak. Insect damage was surveyed using a 4 mm 4 mm grid placed over randomly selected foliage from a subset of samples for each plant species. Foliage at each point on the grid was scored as damaged or undamaged. Damage to pine and fern was consistently small, but more variable for oak. Since damage to some oak trees was severe enough to significantly decrease leaf area and weight, we calculated oak insect densities per leaf to more accurately reflect the tree area sampled. For analysis of richness and diversity, which required whole number data, insect abundance was calculated as insects/50 g fresh weight foliage for fern, insects/50 leaves for oak and insects/100 g fresh weight foliage for pine. These were based on average sample sizes for each plant species.
2.3. Insect community analysis Species richness and Shannon-Weiner diversity (H0 ; Magurran, 1988) were calculated using EcoSim (Gotelli and Entsminger, 1997). Richness estimates
Fig. 1. Map of the San Bernardino National Forest. The study sites (circled) were established in the early 1970s for long-term air pollution monitoring projects by the USDA Forest Service, Pacific Southwest Research Station. There is a steep east/west gradient for ozone and nitrogen deposition. High pollution sites are indicated by black-filled circles, low pollution sites are indicated by white-filled circles.
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herbivores communities, herbivore data from these trees were excluded from all subsequent analyses (with all trees included BP richness in 2001 was 3 1.8 and 2002 was 1 0.8; diversity in 2001 was 0.3 0.3 and 2002 was 0.1 0.2). This made richness and diversity from impacted sites more comparable, but did not significantly change the relationship among sites shown in the discriminant function analysis
were calculated based on rarefaction curves generated by EcoSim, with 100 iterations of Monte Carlo-style simulations from the pooled herbivore communities. Each year, herbivore morphotaxa were pooled within each site and plant species for richness and diversity estimates. This minimized the effects due to clumped species distributions and the small number of samples. For each plant species, total numbers of herbivores were calculated for each site. The site with the smallest number of herbivores collected was used to determine the herbivore abundance level on the rarefaction curves for comparison among sites. Herbivore richness and diversity were compared among sites at the level that corresponded to 80% of the abundance for the site with the lowest herbivore abundance. This standardized sampling intensity among sites. EcoSim uses the mean and variance from Monte Carlo iterations to calculate 95% confidence intervals for species richness and diversity. Sites had significantly different herbivore richness and diversity if their confidence intervals did not overlap. Changes in species composition among sites were evaluated using discriminant function analysis. For this analysis, which required a smaller number of herbivore variables, the most abundant herbivore morphotaxa for each plant species were represented individually while less abundant morphotaxa were grouped by family or feeding guild. This also ensured normal distributions and equal variances of herbivore groups. The discriminant function equation on each axis is a combination of all herbivore variables. Discriminant function loadings (range 1 to 1) were examined to determine which herbivore groups contributed strongly to site separation. High abundance of insects with negative loadings contributed to low discriminant function scores, while high abundance of insects with positive loadings contributed to high discriminant scores. Discriminant function scores were calculated for each tree and each fern sample. Site means and S.E. for discriminant functions were used to evaluate differences in herbivore community structure among the sites. Eigenvalues (l) indicated the strength of the discriminant function. A strong discriminant function with significant separation of sites is indicated when l > 1. Differences in site scores on each discriminant function axis were tested by ANOVA and Fishers LSD at a ¼ 0.05 using SAS (2001).
3.2. Herbivore morphospecies distribution We saw an average of 22 2 morphospecies on fern each year, with an average of 10 morphospecies occurring at each site. Morphospecies unique to one site comprised 2% or less of the total individuals present in fern samples while herbivore morphospecies present at all six sites accounted for 95% of individuals present. The six morphospecies that accounted for the majority of individuals present included thrips, mites, nectar-feeding ants, aphids, a stem-boring Lepidoptera and a foliage feeding sawfly. We collected an average of 43 1 herbivore taxa collected from oak each year and an average of 15 taxa at each site. The unique taxa contributed 4% of the total number of individuals collected. Those taxa (5) that were present at all six sites made up 74% of individuals collected. Individuals collected at 5 or 6 sites (9 taxa) made up 95% of the total number of individuals. These taxa included one free-feeding Lepidoptera, three leaf-tying Lepidoptera, one mirid, an aphid, a gall-forming aphid, a membracid and a thrips. The total number of pine herbivore morphotaxa averaged 32 2 taxa each year. Pine herbivore abundance decreased from 2001 to 2002 by 60% and the average number of species observed at each site decreased by 40%. Those herbivore taxa that were unique to one or two sites accounted for 4% of individuals in 2001 and 18% of individuals in 2002. Individuals from taxa present at all sites accounted for approximately 25% of all individuals in both years. Individuals from taxa present at five or six of the sites accounted for about 70% of all individuals each year. Overall, fern and oak communities were dominated both years by commonly occurring taxa. The contribution of unique taxa to fern and oak communities was consistently small. The contribution of unique taxa to pine communities was more variable, reaching as high as 50% of all individuals at some sites in 2002. Pine communities also had greater turnover in which
3. Results 3.1. Herbivore abundance Generally, total herbivore abundance was similar among sites (Table 2). There were two exceptions that impacted richness and diversity estimates. GV showed very high oak herbivore abundance in 2002 due to very high densities of one morphospecies of aphid on two trees out of ten. We excluded herbivore data from those two trees with high aphid abundance for richness, diversity and the discriminant function analysis (with all trees included GV richness 3 1.8; diversity 0.5 0.2). Pine herbivore abundance was higher at BP compared to all other sites. High pine herbivore abundance was due to one tree in 2001 and two trees 2002 that were infested with high numbers of scale. For site comparison among pine
Table 2 Mean sample abundance (S.E.) for total herbivore abundance for each plant species Fern 2001 BF GV BL CP BP DW
14 28 25 9 16 29
ab
(3) (5)cd (4)bcd (1)a (3)abc (8)d
Fern 2002 264 157 23 25 44 351
cd
(70) (40)bc (4)a (6)a (7)ab (79)d
Oak 2001 21 75 31 97 158 130
a
(3) (10)abc (7)ab (17)bcd (59)d (20)cd
Oak 2002
Pine 2001
Pine 2002
9 1892 23 115 95 125
5 20 25 16 126 4
4 2 10 2 40 1
(3) (1523) (4) (35) (17) (84)
(1) (7) (10) (3) (91) (2)
(0.9)a (0.5)a (2)a (0.8)a (23)b (0.6)a
Herbivore abundance was calculated as insects/50 g foliage for fern, insects/50 leaves for oak and insects/100g foliage for pine, based on average sample sizes for each plant. Means followed by different letters are significantly different at P < 0.05 by ANOVA, fishers LSD (df 5,54).
M. Eatough Jones, T.D. Paine / Environmental Pollution 143 (2006) 377e387
taxa were present at each site from year to year (80%). In contrast, such taxa in fern herbivore communities never made up more than 3% of individuals at one site. The suite of herbivores found on fern and oak was much more consistent between years than for pine. 3.3. Species richness and diversity Fern herbivore morphospecies richness was not significantly different among the sites in 2001 (Table 3). In 2002, fern herbivore richness was higher at the two high ozone sites (CP and BP), but that difference was only significant for the high ozone site CP compared to the two low pollution sites, BF and BL. The largest differences in diversity for fern herbivores were associated with three sites heavily impacted by aphids in 2002. Aphid abundance increased by an order of magnitude from 2001 to 2002 at these sites (BF, GV and DW) which had significantly lower diversity in 2002 than in 2001 (Table 3). Overall comparisons between western sites (CP, BP and DW) which had higher ambient nitrogenous pollution exposure and eastern sites (BF, GV and BL), and comparisons between high ozone sites (CP and BP) and other sites (DW, BF, GV and BL) showed no differences in fern herbivore diversity associated with the air pollution gradient. There were significant differences in oak herbivore richness among sites exposed to low nitrogenous pollution (BF, GV and BL), but not among high nitrogen pollution sites, high ozone sites, or in comparisons between high pollution and low pollution sites (Table 3). There were no consistent differences in oak herbivore diversity between high (CP, BP and DW) and low (BF, GV and BL) nitrogenous pollution sites nor between high ozone sites (CP and BP) and the other sites in either year (Table 3).
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The high nitrogen pollution sites, and particularly the high ozone sites, CP and BP tended to have lower pine herbivore richness than the low pollution sites, but this difference was generally not significant (Table 3). Pine herbivore richness generally decreased within sites from 2001 to 2002. This decrease was largest for the high nitrogen pollution sites CP and DW. Pine herbivore diversity showed similar patterns. At most sites, diversity did not decrease between years as richness did. However diversity did decrease significantly for the high nitrogen pollution sites CP and DW. The high ozone sites CP and BP tended to have lower diversity than low ozone sites in 2002, but this difference was not significant. 3.4. Discriminant function analysis Discriminant function analysis of bracken fern herbivore groups for 2001 yielded one strong discriminant function (l ¼ 2.0, x axis) that contributed to site separation (Fig. 2). Both the nitrogen contrast and the ozone contrast indicated that high and low pollution sites were significantly different. All the high nitrogen pollution sites were significantly different from two low pollution sites, GV and BL, along the first axis (Table 4). BF, a low pollution site, was grouped with the high pollution sites. High pollution sites were associated with higher abundance of sawflies, while low pollution sites were associated with higher abundance of aphids and mites (Table 5, Fig. 2). In 2002, the low pollution sites were significantly different from high nitrogen pollution sites, but not the high ozone sites on the first axis (l ¼ 2.8) (Table 4). The direction of the first discriminant function axis was flipped between 2001 and 2002. However, the spatial relationship among sites was very similar between years, except one high pollution site, CP, was shifted toward the low pollution
Table 3 Herbivore morphotaxa richness and Shannon-Weiner diversity with 95% confidence intervals determined at 80% abundance for the site with the smallest abundance Site
Richness
Diversity
2001 Fern
BL GV BF DW BP CP
8 2.6 9 2.6 9 1.8 8 2.6 7 2.1 8 0.9
Oak
BL GV BF DW BP CP
6 1.4 10 2.9 14 2.0 12 3.4 9 2.7 10 3.3
Pine
BL GV BF DW BP CP
7 3.0 8 3.2 9 1.9 11 1.7 5 2.5 10 2.5
2002
2001
2002
5 0.0 7 2.8 6 1.8 6 2.2 10 1.9 10 2.4
b ab b ab ab a
1.2 0.3 1.5 0.2 1.8 0.1 1.2 0.2 0.8 0.2 1.5 0.1
bc ab a bc c b
1.3 0.0 0.4 0.1 0.7 0.1 0.7 0.1 0.9 0.1 1.4 0.1
a c b b b a
c abc a ab bc abc
6 2.6 6 2.7 11 1.3 8 2.7 9 2.5 9 2.9
b b a ab ab ab
1.6 0.1 1.4 0.1 1.4 0.2 1.2 0.2 1.5 0.2 1.3 0.1
a ab ab b ab b
1.1 0.2 0.8 0.3 1.9 0.1 1.2 0.3 1.7 0.2 0.8 0.3
c c a bc ab c
ab ab ab a b ab
5 1.9 6 1.9 5 1.9 5 0.8 3 1.8 3 1.0
ab a ab b ab b
1.4 0.4 1.5 0.5 1.9 0.2 1.8 0.2 0.9 0.3 1.9 0.3
ab ab a a b a
1.5 0.5 1.7 0.3 1.5 0.5 1.3 0.2 0.8 0.6 0.7 0.3
ab a ab a ab b
Within years, values with different letters are significantly different as determined by the confidence intervals.
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low pollution and high pollution sites (Table 4). Low nitrogen pollution sites were tightly clustered and not significantly different, indicating very similar herbivore communities. The high pollution sites were all significantly different from each other, but all diverged in a similar direction from the low pollution sites. Discriminant function scores on the first axis for the high ozone sites were also significantly different from all other sites in 2001 and 2002. High pollution sites scores on discriminant function 1 were strongly associated with plutellids in 2001. Other Lepidoptera groups contributed more moderately to the positive discriminant function scores associated with the high pollution sites (Table 5). In 2002, plutellids
sites in 2002. The high nitrogen pollution sites, DW and BP, and the low pollution site BF had significantly higher scores on the first axis compared to all other sites (Fig. 2, Table 4). Sawfly abundance was strongly associated with high discriminant function scores and the high pollution sites. Aphids and ants also contributed to high discriminant function scores while mites remained associated with the low pollution sites. The oak herbivore discriminant function analysis showed a very strong separation of high nitrogen pollution and low nitrogen pollution sites in both 2001 and 2002 (Fig. 3). The first axis, with a very strong discriminant function (2001 l ¼ 6.6; 2002 l ¼ 2.5) showed a large and significant separation of
aphids, mites
sawflies 2.5
a
2
GV 1.5
DW 1 0.5
-2
-1.5
-1
0
-0.5
BF
BP
0
0.5
1
1.5
2
2.5
3
-0.5
CP
BL
-1 -1.5
-2 3
b
CP 2
1
BL
BP DW
-3
-2
-1
0
0
1
2
3
BF -1
GV
-2
-3
-4
mites
aphids, ants, sawflies
Fig. 2. Discriminant function scores (site means and S.E.) for the 2001 (a) and 2002 (b) bracken fern herbivore community. Open markers indicate low pollution sites, filled markers indicate high pollution sites. Insect groups on the axes contributed strongly to site separation.
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Table 4 Analysis of variance of discriminant function scores on the first axis for herbivore communities from the six sites along the air pollution gradient
Fern 2001 Fern 2002 Oak 2001 Oak 2002 Pine 2001 Pine 2002
Df
Sites
Nitrogen contrast
Ozone contrast
F
P
F
P
F
P
5, 54 5, 54 5,54 5,52 5,53 5,52
21.4 30.4 71.3 26.2 9.5 14.9
<0.001 <0.001 <0.001 <0.001 <0.001 <0.001
64.8 22.5 234.4 100.8 9.9 30.9
<0.001 <0.001 <0.001 <0.001 0.003 <0.001
25.2 0.23 284.6 57.2 24.4 6.0
<0.001 0.63 <0.001 <0.001 <0.001 0.018
Low pollution sites BF, GV and BL were contrasted to high pollution sites CP, BP and DW for the nitrogen contrast. High ozone sites CP and BP were contrasted to BF, GV, BL and DW for the ozone contrast.
and fruittree leafrollers (Archips argyrospilus) both strongly contributed to high positive scores on the first axis for high pollution sites (Fig. 3). Discriminant function 2 also contributed to site separation in 2001 (l ¼ 1.2). The negative scores on discriminant function 2 for DW and CP were associated with high abundance of the fruittree leafroller (loading ¼ 0.89). In 2002, the low scores on discriminant function 2 for the high ozone sites CP and BP were associated with other species of leaf-tying Lepidoptera (l ¼ 0.8; loading ¼ 0.66). Pine discriminant functions were not as strong as those for oak and fern. Although both the nitrogen contrast and the ozone contrast indicated differences between high and low pollution sites in 2001, there was no pattern among sites associated with the pollution gradient, and the first discriminant function was not strong (l ¼ 0.9) (Table 4, Fig. 4). In 2002, the first axis contributed to separating high and low pollution sites (l ¼ 1.4). This difference was driven by the herbivore community at the low pollution site BL, which had higher scores that were associated with leafhoppers, aphids and adelgids (Table 5). 4. Discussion The discriminant function analysis showed the most consistent patterns of site association along the air pollution gradient
based on insect herbivore communities. The average difference between the low and high nitrogen pollution site means indicated by discriminant function analysis was ten times that indicated by richness and diversity estimates. Oak herbivore communities had the most consistent changes across the air pollution gradient and may provide the best indicators of air pollution effects in the mixed conifer forests of the western United States. Oak herbivore communities at high pollution sites were dominated by chewing insects, particularly leaf-tying and leaf-rolling Lepidoptera. Changes in plant nutrients associated with air pollution exposure could potentially affect abundance of oak lepidopterans (Kyto et al., 1996; Heliovaara and Vaisanen, 1993; Mattson, 1980). However, the effects of ozone, along with nitrogen, may be important for the differences in the response of lepidopterans and sucking insects along the gradient. Many studies have found that other chewing insects showed increased feeding and increased growth rates on ozone fumigated foliage (Whittaker et al., 1990; Coleman and Jones, 1988; Jeffords and Endress, 1984). If air pollution exposure mediated a shift from herbivore communities dominated by sucking insects to communities dominated by chewing insects, it could have an impact on nutrient cycling in these forests. Chewing insect damage may increase litterfall and nutrient leaching from foliage (Kimmins, 1972). Schowalter et al. (1991) found that manipulating levels of sap-sucking herbivores had little effect on nutrient turnover
Table 5 Herbivore community loadings for the first discriminant function axis. Herbivore groups with loadings greater than 0.5 and less than 0.5 contributed strongly to site separation Fern
Oak
Pine
Insect group
2001
2002
Insect group
2001
2002
Insect group
2001
2002
Sawflies Stem-boring Lep. Other chewing Ants Aphids Leafhoppers Other sucking Thrips Mites
0.47 0.31 0.26 0.10 0.58 0.10 0.10 0.14 0.73
0.76 0.29 0.29 0.49 0.46 0.07 0.04 0.09 0.64
Plutellids Fruittree leafroller Tortricid #2 Other rollers Free-feeding Lep. Aphids Gall aphids Leafhoppers Thrips Mirids
0.88 0.16 0.44 0.36 0.43 0.05 0.41 0.32 0.48 0.32
0.85 0.69 0.37 0.48 0.43 0.05 0.20 0.14 0.21 0.46
Sawflies and Lep. Beetles Aphids Adelgids Leafhoppers Scale Mirid #1 Other mirids Thrips Mites
0.56 0.30 0.17 0.56 0.02 0.61 0.13 0.22 0.02 0.01
0.19 0.27 0.65 0.71 0.65 0.02 0.48 0.12 0.30 0.33
0.56
0.40
0.69
0.63
0.41
0.85
Proportion variation
The proportion variation is the amount of variation that contributed to differences between sites explained by the first discriminant function axis.
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plutellids 2
a
1.5
GV
1
BF
0.5
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and litter decomposition, but increased defoliation due to chewing herbivores significantly increased the amount of light and water that reached forest floors. Nutrient turnover rates for nitrogen, potassium and calcium also increased, and litter microarthopod community composition changed (Schowalter and Sabin, 1991). If similar processes operate along the air pollution gradient in the San Bernardino Mountains, increased nutrient flux may alter litter community processes, soil nutrient availability and plant competition, contributing to changes in forest ecosystems impacted by air pollution.
Fern herbivores showed the most fluctuation in herbivore community comparisons within and among low and high pollution sites. The low pollution site, BF, was grouped with high pollution sites, while the high pollution site CP had very different herbivore community characteristics from year to year. One factor that may have influenced the change in site associations between years for fern communities was the phenology of sawflies at CP. Sawfly eggs were present on fern foliage from CP during collection in 2002. Very few eggs were observed on fern in 2001 at any site, although
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Aphids, leafhoppers, adelgids Fig. 4. Discriminant function scores (site means and S.E.) for the 2001 (a) and 2002 (b) pine herbivore community. Open markers indicate low pollution sites, filled markers indicate high pollution sites. Insect groups on the axes contributed strongly to site separation.
fern was collected at all sites at the same apparent stage of foliage expansion (unpublished data). While this study suggests fern herbivores may be affected by air pollution, other factors also likely play a strong roll in fern herbivore community structure, making it less effective for monitoring impacts. Changes in pine herbivore communities where not strongly associated with the air pollution gradient when several sites across the gradient were examined. Because no consistent pattern of difference was found in herbivore communities when several sites were compared, it is unlikely that differences between these sites were due to differences in air pollution exposure despite the fact that previous research has clearly shown that ponderosa pine in the San Bernardino Mountains have been strongly affected by both ozone exposure and nitrogen deposition (Grulke and Balduman, 1999; Miller and Rechel,
1999; Fenn et al., 1996). Increased nitrogen availability may not increase host suitability of conifers for foliage-feeding insects as it does for other plants (Kyto et al., 1996). Although other factors also vary along the gradient in the San Bernardino Mountains, our data suggest that air pollution exposure may play a role in altering community structure. In the two years of this study we saw patterns of change among sites that were strongly associated with the air pollution gradient. Several studies that have shown large changes in herbivore communities sampled over an elevation gradient with changing vegetation regimes (e.g. Ribeiro et al., 1998; Hagvar, 1976; Janzen et al., 1976). Our study sites were contained within the mixed conifer zone to avoid such effects. Even within a vegetation zone, increasing elevation is usually associated with decreased arthropod abundance and decreased
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species richness (e.g. Fernandes et al., 2004; Reynolds and Crossley, 1997). We did not see decreased abundance or richness associated with changes either in elevation or pollution exposure. Dominant herbivores remained the same for all sites across the gradient, but associations of characteristic herbivores changed between high and low pollution sites. Additionally, five decades of air pollution research in the San Bernardino Mountains have confirmed many air pollution impacts already affecting soil processes and plant growth (Miller and McBride, 1999; Bytnerowicz and Fenn, 1996; Fenn et al., 1996). Oak herbivores showed strong differences along the pollution gradient with communities at high pollution sites dominated by chewing insects but there were no significant differences among sucking insects along the gradient. To conclude, while diversity and richness have traditionally been used to monitor environmental impacts on ecological communities, had we relied on these measures, we would have detected no effect of air pollution on herbivore communities along the air pollution gradient in the San Bernardino Mountains. In spite of uniform herbivore richness and diversity across the gradient, there were shifts in herbivore community organization that were strongly associated with air pollution exposure, and that have the potential to affect forest processes. These shifts in herbivore communities could affect resources for higher trophic levels as well as influence rates of nutrient cycling. Monitoring changes in oak herbivore communities through discriminant function analysis may serve as an important indicator for assessing the effects of air pollution in xeric western forests.
Acknowledgements We are grateful to Stephanie Washburn, Chris Hanlon, Stuart Wooley, Sarah Allen, Idara Essien, and Jocelyn Holt for their assistance in the field and laboratory. We thank J. Daniel Hare, Richard Redak, Edith Allen, Rocio Alonso and Mark Fenn for their comments on the manuscript. Funding was provided in part by the University of California Graduate Division through a dissertation improvement grant and by a grant from the University of California Academic Senate.
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Detecting changes in insect herbivore communities along a pollution gradient Michele Eatough Jones*, Timothy D. Paine Department of Entomology, University of California Riverside, Riverside, CA 92521, USA Received 13 September 2005; received in revised form 10 December 2005; accepted 14 December 2005
Differences in insect herbivore communities were associated with an ambient air pollution gradient in the mixed conifer forest outside the Los Angeles area. Abstract The forests surrounding the urban areas of the Los Angeles basin are impacted by ozone and nitrogen pollutants arising from urban areas. We examined changes in the herbivore communities of three prominent plant species (ponderosa pine, California black oak and bracken fern) at six sites along an air pollution gradient. Insects were extracted from foliage samples collected in spring, as foliage reached full expansion. Community differences were evaluated using total herbivore abundance, richness, Shannon-Weiner diversity, and discriminant function analysis. Even without conspicuous changes in total numbers, diversity or richness of herbivores, herbivore groups showed patterns of change that followed the air pollution gradient that were apparent through discriminant function analysis. For bracken fern and oak, chewing insects were more dominant at high pollution sites. Oak herbivore communities showed the strongest effect. These changes in herbivore communities may affect nutrient cycling in forest systems. Ó 2005 Elsevier Ltd. All rights reserved. Keywords: Nitrogen deposition; Pinus ponderosa; Quercus kelloggii; Pteridium aquilinum; San Bernardino Mountains
1. Introduction Human activities are widely reported to impact ecosystems on both local and global scales. Long term impacts can include changes in biogeochemical cycles, primary productivity, altered biotic interactions and decreased ecosystem resilience (Chapin et al., 2000; Naeem and Li, 1997; Taylor et al., 1994). Loss of species is assumed to be one of the driving mechanisms for changes in ecosystem function (Chapin et al., 2000). Monitoring species richness or individual species abundance has been promoted as a key method for detecting environmental impacts in many conservation and management programs (USDA Forest Service, 2002).
* Corresponding author. Tel.: þ1 951 827 4488; fax: þ1 951 827 3086. E-mail addresses: [email protected] (M. Eatough Jones), timothy. [email protected] (T.D. Paine). 0269-7491/$ - see front matter Ó 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2005.12.013
Insect herbivore communities may be good indicators of ecosystem changes. Insect herbivore populations can respond to changes in plant nutrients (Mattson, 1980), plant defensive chemistry (Rosenthal and Berenbaum, 1992; Feeny, 1976), plant growth characteristics (Mopper and Simberloff, 1995), or plant community structure (Jefferies and Maron, 1997; Brown et al., 1988). Insect herbivores also play critical roles in ecosystem processes. They influence nutrient cycling (Reynolds and Hunter, 2001; Belovsky and Slade, 2000; Schowalter, 1995) and provide food for higher trophic levels. Changes in the population density or structure of the herbivorous insect community can affect plant productivity, competition between plants, and plant community composition (Ritchie et al., 1998; Jefferies and Maron, 1997; Janzen, 1987). Major anthropogenic impacts affecting forest systems include habitat fragmentation, invasive introduced species, global warming, increased atmospheric CO2, and air pollution.
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Although air pollution affects many ecosystem types worldwide, forests in particular may be vulnerable to air pollution effects (Fowler et al., 1999). Ozone and nitrogen deposition are considered to be the two most important pollutants affecting forest ecosystems in North America (Fowler et al., 1999; Taylor et al., 1994). Atmospheric nitrogen deposition can acidify forest soils and promote leaching of base cations which may cause nutrient imbalances in plants that contribute to forest decline (McNulty and Aber, 1993). Nitric acid deposited on foliar surfaces may directly damage the plant cuticle resulting in increased loss of water and metabolites and greater susceptibility to insects or pathogens (Bytnerowicz et al., 2001). Deposition of atmospheric nitrogen can also act as a fertilizer, stimulating forest growth and increasing plant nitrogen content (Vitousek et al., 1997). Elevated nitrogen in plant tissues has been associated with increased insect performance and greater damage to plants (Kyto et al., 1996; Waring and Cobb, 1992; Lightfoot and Whitford, 1990; Skeffington and Wilson, 1988; Bryant et al., 1983; Mattson, 1980). Alternatively, higher nitrogen in plant tissues may lead to deterrence of insect herbivores through mechanisms including increased nitrate in plant tissues (Jansson and Smilowitz, 1985; Manglitz et al., 1976) and increased resin content in pines (Kyto et al., 1996). Ozone damaged plants may have higher concentrations of free amino acids, soluble proteins and soluble sugars that increase plant nutrients available for insects (Tingey et al., 1976; Barnes, 1972). Most chewing insects studied showed increased feeding and faster growth rates on ozone fumigated foliage (Heliovaara and Vaisanen, 1993). However, results for aphid feeding and performance were more varied. Other factors may also mitigate herbivore responses to ozone-stressed plants. In long-term population surveys, Percy et al. (2002) found no change in aphid performance on ozone fumigated trees, but aphid abundance increased, presumably through the negative effects of ozone on natural enemies of aphids. 1.1. Air pollution in southern California forests Atmospheric pollutants, particularly nitrogenous compounds and ozone are transported inland from urban areas of the Los Angeles basin, along the western slopes of the San Bernardino Mountains by prevailing weather patterns (Miller, 1992). Concern with air pollution damage in the San Bernardino Forest began as early as the 1960s when ozone injury on pine needle tissue was observed (Cobb and Stark, 1970; Miller et al., 1963). The wildland ecosystems surrounding Los Angeles receive the highest levels of nitrogen deposition recorded in North America (Fenn et al., 2003a). Concentrations of tropospheric ozone are also high in this region (Grulke and Balduman, 1999). Chronic exposure to air pollutants has affected plant growth and other plant characteristics (Grulke, 1999; Grulke et al., 1998; Miller and Rechel, 1999; Bytnerowicz and Fenn, 1996). This may potentially affect performance of individual plant species, competitive relationships within the plant community, and the structure and abundance of organisms in other trophic levels (Fenn et al., 2003b).
A well-established horizontal pollution gradient has been described in the San Bernardino Mountains with decreasing ozone and nitrogenous pollutant concentrations from west to east (Miller and Rechel, 1999; Miller et al., 1986). Air pollution impacts to plants have been established along this gradient. Ponderosa pine (Pinus ponderosa Laws) trees from highly polluted western areas have shown greater symptoms associated with ozone exposure than trees from less impacted eastern areas, including decreased growth (Miller, 1992), reduced root mass and belowground allocation of carbohydrates (Grulke et al., 1998), and chlorotic mottling of needles and reduced needle retention (Miller, 1973). Ponderosa pine is among the most susceptible plant species to ozone exposure in the San Bernardino Mountains (Miller et al., 1983), and many have already died, most likely due to a combination of stresses, including oxidant injury, bark beetle attacks and disease (Pronos et al., 1999). Total foliar nitrogen and nitrate were higher in bracken fern (Pteridium aquilinum var. pubescens Underw.), California black oak (Quercus kelloggii Newb.) and ponderosa pine from a western site compared to an eastern site (Fenn et al., 1996). We examined arthropod herbivore communities of three prominent plant species along the ambient pollution gradient in the San Bernardino Mountains. Although it is not possible to rule out effects from other environmental factors in any ambient gradient study, the strong history of documented air pollution effects along this gradient make it an ideal location to examine insect community impacts. We hypothesized that high atmospheric inputs of ozone and nitrogen deposition would lead to alterations in the composition of the insect herbivore communities. To assess if there were changes in herbivore communities along the gradient we used the traditional measures of richness and diversity, and also a multivariate analysis to examine community structure. We evaluated the performance of these different methods for examining herbivore communities and which methods of community comparison gave the clearest and most reproducible results from year to year. 2. Methods 2.1. Study locations Study sites were selected from permanent plots established in the early 1970s by the USDA Forest Service Pacific Southwest Research Station for long-term air pollution monitoring (Arbaugh et al., 2003). Sites were chosen along the west-east air pollution gradient in the mixed conifer zone (1300 m to 2700 m elevation) of the San Bernardino Mountains. Sites selected for this study range from 1500 m to 2300 m (Table 1). The dominant overstory species are ponderosa pine, California black oak and incense cedar (Calocedrus decurrens (Torr.) Florin.) with a minor component of sugar pine (Pinus lambertiana Dougl.) and white fir (Abies concolor Gord. & Glend.). Jeffrey pine (Pinus jeffreyi Grev. & Balf.) becomes more common at higher elevations, while incense cedar becomes less common. Bracken fern forms a dense understory in many locations. Additional details on climate, soils and vegetation are reported in Miller and McBride (1999). Of our study species, ponderosa pine is the most sensitive to ozone exposure (Miller and Rechel, 1999), while bracken fern has shown the greatest increase in foliar nitrogen content along the gradient (Fenn et al., 1996). All sites were subjected to drought conditions during this study. Annual precipitation in the San Bernardino
M. Eatough Jones, T.D. Paine / Environmental Pollution 143 (2006) 377e387 Table 1 Summer air pollution concentrations and geographical characteristics for sites in the San Bernardino Mountains Relative Elev. Distance O3 a NO2a HNO3a pollution (m) from L.A. (ppb) (ppb) (mg/m3) exposure (km) BL GV
Bluff Lake Green Valley Creek BF Barton Flats DW Dogwood BP Breezy Point CP Camp Paivika
Low Low
2320 117 1950 106
59.3
1.4
2.6
Low High High High
1900 125 1725 94.5 1525 86.1 1600 84.7
55.9 59.4 63.2 62.8
3.1 7.7 5.4 6.9
2.0 3.7 5.3 4.8
Air pollution data was not collected at BL during the study period but exposure is historically low. a Average atmospheric concentration determined by passive monitoring (Alonso et al., 2002; Bytnerowicz et al., 2002), June 7 through August 14, 2001. (Atmospheric Deposition Group, USDA PSW Research Station).
Mountains was below half-normal for four years beginning in 1999 and continuing through the study period (California Department of Water Resources, public information). Three western sites had high pollution inputs: Camp Paivika (CP), Breezy Point (BP) and Dogwood (DW). More eastern sites at Barton Flats (BF), Green Valley Creek (GV) and Bluff Lake (BL) had lower atmospheric inputs (Fig. 1). Elevation and precipitation also vary along the gradient, but pines along the gradient showed a strong east to west trend for increased foliar injury indicated by forest pest management index scores (Miller and Rechel, 1999). Recent air pollution concentrations suggested two contrasts among sites (Table 1). Data on nitrogen oxides suggested categorizing BF, GV and BL as low pollution and CP, BP and DW as high pollution sites. However, ozone concentrations at DW were more similar to those recorded at the low pollution sites. Ozone data suggested a contrast between CP and BP as high exposure sites and DW, BF, GV and BL as low pollution sites. We evaluated both these contrasts in our analysis of herbivore community characteristics. Contrasts were designed using the study sites as the unit of replication, with herbivore
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communities analyzed separately each year to mitigate lack of independence of replicates from resampling in the same locations each year.
2.2. Sample collection Insect and foliage collections from oak, pine and fern were taken as foliage reached full expansion in 2001 and 2002. Collection times were late May into June for fern and oak, and July for pine. Foliage samples of three 30 cm long branch tips were collected at mid-crown height from 10 oak and 10 pine at each site. Pine foliage included current year’s needles and older annual whorls on the branch. Foliage was removed using a pole-pruner with a basket lined with a 125 l trash bag attached below the pruner. Foliage was collected in a bag so that the bag could be closed immediately after collection to minimize loss of insect fauna through handling the sample. The above ground growth of bracken fern was bagged and removed from ten 0.5 m 0.5 m areas randomly selected at each study site. Collected foliage was stored on ice for transport and stored in an ultracold (65 C) freezer in the laboratory. Insects were manually removed from the foliage, and identified to family and morphotaxa. Morphotaxa were assigned based on visually distinct appearances. Common morphotaxa were collected from foliage near plot areas and reared in the lab so that all stages could be identified and grouped. Herbivore abundance was calculated as insects/gram fresh weight foliage for pine and fern and as insects/leaf for oak. Insect damage was surveyed using a 4 mm 4 mm grid placed over randomly selected foliage from a subset of samples for each plant species. Foliage at each point on the grid was scored as damaged or undamaged. Damage to pine and fern was consistently small, but more variable for oak. Since damage to some oak trees was severe enough to significantly decrease leaf area and weight, we calculated oak insect densities per leaf to more accurately reflect the tree area sampled. For analysis of richness and diversity, which required whole number data, insect abundance was calculated as insects/50 g fresh weight foliage for fern, insects/50 leaves for oak and insects/100 g fresh weight foliage for pine. These were based on average sample sizes for each plant species.
2.3. Insect community analysis Species richness and Shannon-Weiner diversity (H0 ; Magurran, 1988) were calculated using EcoSim (Gotelli and Entsminger, 1997). Richness estimates
Fig. 1. Map of the San Bernardino National Forest. The study sites (circled) were established in the early 1970s for long-term air pollution monitoring projects by the USDA Forest Service, Pacific Southwest Research Station. There is a steep east/west gradient for ozone and nitrogen deposition. High pollution sites are indicated by black-filled circles, low pollution sites are indicated by white-filled circles.
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herbivores communities, herbivore data from these trees were excluded from all subsequent analyses (with all trees included BP richness in 2001 was 3 1.8 and 2002 was 1 0.8; diversity in 2001 was 0.3 0.3 and 2002 was 0.1 0.2). This made richness and diversity from impacted sites more comparable, but did not significantly change the relationship among sites shown in the discriminant function analysis
were calculated based on rarefaction curves generated by EcoSim, with 100 iterations of Monte Carlo-style simulations from the pooled herbivore communities. Each year, herbivore morphotaxa were pooled within each site and plant species for richness and diversity estimates. This minimized the effects due to clumped species distributions and the small number of samples. For each plant species, total numbers of herbivores were calculated for each site. The site with the smallest number of herbivores collected was used to determine the herbivore abundance level on the rarefaction curves for comparison among sites. Herbivore richness and diversity were compared among sites at the level that corresponded to 80% of the abundance for the site with the lowest herbivore abundance. This standardized sampling intensity among sites. EcoSim uses the mean and variance from Monte Carlo iterations to calculate 95% confidence intervals for species richness and diversity. Sites had significantly different herbivore richness and diversity if their confidence intervals did not overlap. Changes in species composition among sites were evaluated using discriminant function analysis. For this analysis, which required a smaller number of herbivore variables, the most abundant herbivore morphotaxa for each plant species were represented individually while less abundant morphotaxa were grouped by family or feeding guild. This also ensured normal distributions and equal variances of herbivore groups. The discriminant function equation on each axis is a combination of all herbivore variables. Discriminant function loadings (range 1 to 1) were examined to determine which herbivore groups contributed strongly to site separation. High abundance of insects with negative loadings contributed to low discriminant function scores, while high abundance of insects with positive loadings contributed to high discriminant scores. Discriminant function scores were calculated for each tree and each fern sample. Site means and S.E. for discriminant functions were used to evaluate differences in herbivore community structure among the sites. Eigenvalues (l) indicated the strength of the discriminant function. A strong discriminant function with significant separation of sites is indicated when l > 1. Differences in site scores on each discriminant function axis were tested by ANOVA and Fishers LSD at a ¼ 0.05 using SAS (2001).
3.2. Herbivore morphospecies distribution We saw an average of 22 2 morphospecies on fern each year, with an average of 10 morphospecies occurring at each site. Morphospecies unique to one site comprised 2% or less of the total individuals present in fern samples while herbivore morphospecies present at all six sites accounted for 95% of individuals present. The six morphospecies that accounted for the majority of individuals present included thrips, mites, nectar-feeding ants, aphids, a stem-boring Lepidoptera and a foliage feeding sawfly. We collected an average of 43 1 herbivore taxa collected from oak each year and an average of 15 taxa at each site. The unique taxa contributed 4% of the total number of individuals collected. Those taxa (5) that were present at all six sites made up 74% of individuals collected. Individuals collected at 5 or 6 sites (9 taxa) made up 95% of the total number of individuals. These taxa included one free-feeding Lepidoptera, three leaf-tying Lepidoptera, one mirid, an aphid, a gall-forming aphid, a membracid and a thrips. The total number of pine herbivore morphotaxa averaged 32 2 taxa each year. Pine herbivore abundance decreased from 2001 to 2002 by 60% and the average number of species observed at each site decreased by 40%. Those herbivore taxa that were unique to one or two sites accounted for 4% of individuals in 2001 and 18% of individuals in 2002. Individuals from taxa present at all sites accounted for approximately 25% of all individuals in both years. Individuals from taxa present at five or six of the sites accounted for about 70% of all individuals each year. Overall, fern and oak communities were dominated both years by commonly occurring taxa. The contribution of unique taxa to fern and oak communities was consistently small. The contribution of unique taxa to pine communities was more variable, reaching as high as 50% of all individuals at some sites in 2002. Pine communities also had greater turnover in which
3. Results 3.1. Herbivore abundance Generally, total herbivore abundance was similar among sites (Table 2). There were two exceptions that impacted richness and diversity estimates. GV showed very high oak herbivore abundance in 2002 due to very high densities of one morphospecies of aphid on two trees out of ten. We excluded herbivore data from those two trees with high aphid abundance for richness, diversity and the discriminant function analysis (with all trees included GV richness 3 1.8; diversity 0.5 0.2). Pine herbivore abundance was higher at BP compared to all other sites. High pine herbivore abundance was due to one tree in 2001 and two trees 2002 that were infested with high numbers of scale. For site comparison among pine
Table 2 Mean sample abundance (S.E.) for total herbivore abundance for each plant species Fern 2001 BF GV BL CP BP DW
14 28 25 9 16 29
ab
(3) (5)cd (4)bcd (1)a (3)abc (8)d
Fern 2002 264 157 23 25 44 351
cd
(70) (40)bc (4)a (6)a (7)ab (79)d
Oak 2001 21 75 31 97 158 130
a
(3) (10)abc (7)ab (17)bcd (59)d (20)cd
Oak 2002
Pine 2001
Pine 2002
9 1892 23 115 95 125
5 20 25 16 126 4
4 2 10 2 40 1
(3) (1523) (4) (35) (17) (84)
(1) (7) (10) (3) (91) (2)
(0.9)a (0.5)a (2)a (0.8)a (23)b (0.6)a
Herbivore abundance was calculated as insects/50 g foliage for fern, insects/50 leaves for oak and insects/100g foliage for pine, based on average sample sizes for each plant. Means followed by different letters are significantly different at P < 0.05 by ANOVA, fishers LSD (df 5,54).
M. Eatough Jones, T.D. Paine / Environmental Pollution 143 (2006) 377e387
taxa were present at each site from year to year (80%). In contrast, such taxa in fern herbivore communities never made up more than 3% of individuals at one site. The suite of herbivores found on fern and oak was much more consistent between years than for pine. 3.3. Species richness and diversity Fern herbivore morphospecies richness was not significantly different among the sites in 2001 (Table 3). In 2002, fern herbivore richness was higher at the two high ozone sites (CP and BP), but that difference was only significant for the high ozone site CP compared to the two low pollution sites, BF and BL. The largest differences in diversity for fern herbivores were associated with three sites heavily impacted by aphids in 2002. Aphid abundance increased by an order of magnitude from 2001 to 2002 at these sites (BF, GV and DW) which had significantly lower diversity in 2002 than in 2001 (Table 3). Overall comparisons between western sites (CP, BP and DW) which had higher ambient nitrogenous pollution exposure and eastern sites (BF, GV and BL), and comparisons between high ozone sites (CP and BP) and other sites (DW, BF, GV and BL) showed no differences in fern herbivore diversity associated with the air pollution gradient. There were significant differences in oak herbivore richness among sites exposed to low nitrogenous pollution (BF, GV and BL), but not among high nitrogen pollution sites, high ozone sites, or in comparisons between high pollution and low pollution sites (Table 3). There were no consistent differences in oak herbivore diversity between high (CP, BP and DW) and low (BF, GV and BL) nitrogenous pollution sites nor between high ozone sites (CP and BP) and the other sites in either year (Table 3).
381
The high nitrogen pollution sites, and particularly the high ozone sites, CP and BP tended to have lower pine herbivore richness than the low pollution sites, but this difference was generally not significant (Table 3). Pine herbivore richness generally decreased within sites from 2001 to 2002. This decrease was largest for the high nitrogen pollution sites CP and DW. Pine herbivore diversity showed similar patterns. At most sites, diversity did not decrease between years as richness did. However diversity did decrease significantly for the high nitrogen pollution sites CP and DW. The high ozone sites CP and BP tended to have lower diversity than low ozone sites in 2002, but this difference was not significant. 3.4. Discriminant function analysis Discriminant function analysis of bracken fern herbivore groups for 2001 yielded one strong discriminant function (l ¼ 2.0, x axis) that contributed to site separation (Fig. 2). Both the nitrogen contrast and the ozone contrast indicated that high and low pollution sites were significantly different. All the high nitrogen pollution sites were significantly different from two low pollution sites, GV and BL, along the first axis (Table 4). BF, a low pollution site, was grouped with the high pollution sites. High pollution sites were associated with higher abundance of sawflies, while low pollution sites were associated with higher abundance of aphids and mites (Table 5, Fig. 2). In 2002, the low pollution sites were significantly different from high nitrogen pollution sites, but not the high ozone sites on the first axis (l ¼ 2.8) (Table 4). The direction of the first discriminant function axis was flipped between 2001 and 2002. However, the spatial relationship among sites was very similar between years, except one high pollution site, CP, was shifted toward the low pollution
Table 3 Herbivore morphotaxa richness and Shannon-Weiner diversity with 95% confidence intervals determined at 80% abundance for the site with the smallest abundance Site
Richness
Diversity
2001 Fern
BL GV BF DW BP CP
8 2.6 9 2.6 9 1.8 8 2.6 7 2.1 8 0.9
Oak
BL GV BF DW BP CP
6 1.4 10 2.9 14 2.0 12 3.4 9 2.7 10 3.3
Pine
BL GV BF DW BP CP
7 3.0 8 3.2 9 1.9 11 1.7 5 2.5 10 2.5
2002
2001
2002
5 0.0 7 2.8 6 1.8 6 2.2 10 1.9 10 2.4
b ab b ab ab a
1.2 0.3 1.5 0.2 1.8 0.1 1.2 0.2 0.8 0.2 1.5 0.1
bc ab a bc c b
1.3 0.0 0.4 0.1 0.7 0.1 0.7 0.1 0.9 0.1 1.4 0.1
a c b b b a
c abc a ab bc abc
6 2.6 6 2.7 11 1.3 8 2.7 9 2.5 9 2.9
b b a ab ab ab
1.6 0.1 1.4 0.1 1.4 0.2 1.2 0.2 1.5 0.2 1.3 0.1
a ab ab b ab b
1.1 0.2 0.8 0.3 1.9 0.1 1.2 0.3 1.7 0.2 0.8 0.3
c c a bc ab c
ab ab ab a b ab
5 1.9 6 1.9 5 1.9 5 0.8 3 1.8 3 1.0
ab a ab b ab b
1.4 0.4 1.5 0.5 1.9 0.2 1.8 0.2 0.9 0.3 1.9 0.3
ab ab a a b a
1.5 0.5 1.7 0.3 1.5 0.5 1.3 0.2 0.8 0.6 0.7 0.3
ab a ab a ab b
Within years, values with different letters are significantly different as determined by the confidence intervals.
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low pollution and high pollution sites (Table 4). Low nitrogen pollution sites were tightly clustered and not significantly different, indicating very similar herbivore communities. The high pollution sites were all significantly different from each other, but all diverged in a similar direction from the low pollution sites. Discriminant function scores on the first axis for the high ozone sites were also significantly different from all other sites in 2001 and 2002. High pollution sites scores on discriminant function 1 were strongly associated with plutellids in 2001. Other Lepidoptera groups contributed more moderately to the positive discriminant function scores associated with the high pollution sites (Table 5). In 2002, plutellids
sites in 2002. The high nitrogen pollution sites, DW and BP, and the low pollution site BF had significantly higher scores on the first axis compared to all other sites (Fig. 2, Table 4). Sawfly abundance was strongly associated with high discriminant function scores and the high pollution sites. Aphids and ants also contributed to high discriminant function scores while mites remained associated with the low pollution sites. The oak herbivore discriminant function analysis showed a very strong separation of high nitrogen pollution and low nitrogen pollution sites in both 2001 and 2002 (Fig. 3). The first axis, with a very strong discriminant function (2001 l ¼ 6.6; 2002 l ¼ 2.5) showed a large and significant separation of
aphids, mites
sawflies 2.5
a
2
GV 1.5
DW 1 0.5
-2
-1.5
-1
0
-0.5
BF
BP
0
0.5
1
1.5
2
2.5
3
-0.5
CP
BL
-1 -1.5
-2 3
b
CP 2
1
BL
BP DW
-3
-2
-1
0
0
1
2
3
BF -1
GV
-2
-3
-4
mites
aphids, ants, sawflies
Fig. 2. Discriminant function scores (site means and S.E.) for the 2001 (a) and 2002 (b) bracken fern herbivore community. Open markers indicate low pollution sites, filled markers indicate high pollution sites. Insect groups on the axes contributed strongly to site separation.
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Table 4 Analysis of variance of discriminant function scores on the first axis for herbivore communities from the six sites along the air pollution gradient
Fern 2001 Fern 2002 Oak 2001 Oak 2002 Pine 2001 Pine 2002
Df
Sites
Nitrogen contrast
Ozone contrast
F
P
F
P
F
P
5, 54 5, 54 5,54 5,52 5,53 5,52
21.4 30.4 71.3 26.2 9.5 14.9
<0.001 <0.001 <0.001 <0.001 <0.001 <0.001
64.8 22.5 234.4 100.8 9.9 30.9
<0.001 <0.001 <0.001 <0.001 0.003 <0.001
25.2 0.23 284.6 57.2 24.4 6.0
<0.001 0.63 <0.001 <0.001 <0.001 0.018
Low pollution sites BF, GV and BL were contrasted to high pollution sites CP, BP and DW for the nitrogen contrast. High ozone sites CP and BP were contrasted to BF, GV, BL and DW for the ozone contrast.
and fruittree leafrollers (Archips argyrospilus) both strongly contributed to high positive scores on the first axis for high pollution sites (Fig. 3). Discriminant function 2 also contributed to site separation in 2001 (l ¼ 1.2). The negative scores on discriminant function 2 for DW and CP were associated with high abundance of the fruittree leafroller (loading ¼ 0.89). In 2002, the low scores on discriminant function 2 for the high ozone sites CP and BP were associated with other species of leaf-tying Lepidoptera (l ¼ 0.8; loading ¼ 0.66). Pine discriminant functions were not as strong as those for oak and fern. Although both the nitrogen contrast and the ozone contrast indicated differences between high and low pollution sites in 2001, there was no pattern among sites associated with the pollution gradient, and the first discriminant function was not strong (l ¼ 0.9) (Table 4, Fig. 4). In 2002, the first axis contributed to separating high and low pollution sites (l ¼ 1.4). This difference was driven by the herbivore community at the low pollution site BL, which had higher scores that were associated with leafhoppers, aphids and adelgids (Table 5). 4. Discussion The discriminant function analysis showed the most consistent patterns of site association along the air pollution gradient
based on insect herbivore communities. The average difference between the low and high nitrogen pollution site means indicated by discriminant function analysis was ten times that indicated by richness and diversity estimates. Oak herbivore communities had the most consistent changes across the air pollution gradient and may provide the best indicators of air pollution effects in the mixed conifer forests of the western United States. Oak herbivore communities at high pollution sites were dominated by chewing insects, particularly leaf-tying and leaf-rolling Lepidoptera. Changes in plant nutrients associated with air pollution exposure could potentially affect abundance of oak lepidopterans (Kyto et al., 1996; Heliovaara and Vaisanen, 1993; Mattson, 1980). However, the effects of ozone, along with nitrogen, may be important for the differences in the response of lepidopterans and sucking insects along the gradient. Many studies have found that other chewing insects showed increased feeding and increased growth rates on ozone fumigated foliage (Whittaker et al., 1990; Coleman and Jones, 1988; Jeffords and Endress, 1984). If air pollution exposure mediated a shift from herbivore communities dominated by sucking insects to communities dominated by chewing insects, it could have an impact on nutrient cycling in these forests. Chewing insect damage may increase litterfall and nutrient leaching from foliage (Kimmins, 1972). Schowalter et al. (1991) found that manipulating levels of sap-sucking herbivores had little effect on nutrient turnover
Table 5 Herbivore community loadings for the first discriminant function axis. Herbivore groups with loadings greater than 0.5 and less than 0.5 contributed strongly to site separation Fern
Oak
Pine
Insect group
2001
2002
Insect group
2001
2002
Insect group
2001
2002
Sawflies Stem-boring Lep. Other chewing Ants Aphids Leafhoppers Other sucking Thrips Mites
0.47 0.31 0.26 0.10 0.58 0.10 0.10 0.14 0.73
0.76 0.29 0.29 0.49 0.46 0.07 0.04 0.09 0.64
Plutellids Fruittree leafroller Tortricid #2 Other rollers Free-feeding Lep. Aphids Gall aphids Leafhoppers Thrips Mirids
0.88 0.16 0.44 0.36 0.43 0.05 0.41 0.32 0.48 0.32
0.85 0.69 0.37 0.48 0.43 0.05 0.20 0.14 0.21 0.46
Sawflies and Lep. Beetles Aphids Adelgids Leafhoppers Scale Mirid #1 Other mirids Thrips Mites
0.56 0.30 0.17 0.56 0.02 0.61 0.13 0.22 0.02 0.01
0.19 0.27 0.65 0.71 0.65 0.02 0.48 0.12 0.30 0.33
0.56
0.40
0.69
0.63
0.41
0.85
Proportion variation
The proportion variation is the amount of variation that contributed to differences between sites explained by the first discriminant function axis.
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plutellids 2
a
1.5
GV
1
BF
0.5
BL -3
BP
-2
-1
0
0
1
2
3
4
5
6
-0.5
CP
-1 -1.5 -2
DW -2.5 -3
b 2
DW
1.5
1
BF
0.5
-2
-1
BL
GV
0
-0.5
0
1
2
3
4
CP BP
-1
-1.5
Fruittree leafroller, plutellids Fig. 3. Discriminant function scores (site means and S.E.) for the 2001 (a) and 2002 (b) oak herbivore community. Low pollution sites are indicated by open markers, high pollution sites are indicated by filled markers. Insect groups on the axes contributed strongly to site separation.
and litter decomposition, but increased defoliation due to chewing herbivores significantly increased the amount of light and water that reached forest floors. Nutrient turnover rates for nitrogen, potassium and calcium also increased, and litter microarthopod community composition changed (Schowalter and Sabin, 1991). If similar processes operate along the air pollution gradient in the San Bernardino Mountains, increased nutrient flux may alter litter community processes, soil nutrient availability and plant competition, contributing to changes in forest ecosystems impacted by air pollution.
Fern herbivores showed the most fluctuation in herbivore community comparisons within and among low and high pollution sites. The low pollution site, BF, was grouped with high pollution sites, while the high pollution site CP had very different herbivore community characteristics from year to year. One factor that may have influenced the change in site associations between years for fern communities was the phenology of sawflies at CP. Sawfly eggs were present on fern foliage from CP during collection in 2002. Very few eggs were observed on fern in 2001 at any site, although
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Lepidoptera and sawflies, adelgids, scale 2
a
1.5
CP 1
GV 0.5
-1.5
-1
0
-0.5
0
0.5
1
1.5
2
BL
3
BP
-0.5
BF
2.5
DW -1
b 2.5 2
BP
1.5 1 0.5
BF -2
-1.5
-1
-0.5
BL 0
0
0.5
1
1.5
2
2.5
3
3.5
DW CP
-0 .5
GV -1
Aphids, leafhoppers, adelgids Fig. 4. Discriminant function scores (site means and S.E.) for the 2001 (a) and 2002 (b) pine herbivore community. Open markers indicate low pollution sites, filled markers indicate high pollution sites. Insect groups on the axes contributed strongly to site separation.
fern was collected at all sites at the same apparent stage of foliage expansion (unpublished data). While this study suggests fern herbivores may be affected by air pollution, other factors also likely play a strong roll in fern herbivore community structure, making it less effective for monitoring impacts. Changes in pine herbivore communities where not strongly associated with the air pollution gradient when several sites across the gradient were examined. Because no consistent pattern of difference was found in herbivore communities when several sites were compared, it is unlikely that differences between these sites were due to differences in air pollution exposure despite the fact that previous research has clearly shown that ponderosa pine in the San Bernardino Mountains have been strongly affected by both ozone exposure and nitrogen deposition (Grulke and Balduman, 1999; Miller and Rechel,
1999; Fenn et al., 1996). Increased nitrogen availability may not increase host suitability of conifers for foliage-feeding insects as it does for other plants (Kyto et al., 1996). Although other factors also vary along the gradient in the San Bernardino Mountains, our data suggest that air pollution exposure may play a role in altering community structure. In the two years of this study we saw patterns of change among sites that were strongly associated with the air pollution gradient. Several studies that have shown large changes in herbivore communities sampled over an elevation gradient with changing vegetation regimes (e.g. Ribeiro et al., 1998; Hagvar, 1976; Janzen et al., 1976). Our study sites were contained within the mixed conifer zone to avoid such effects. Even within a vegetation zone, increasing elevation is usually associated with decreased arthropod abundance and decreased
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species richness (e.g. Fernandes et al., 2004; Reynolds and Crossley, 1997). We did not see decreased abundance or richness associated with changes either in elevation or pollution exposure. Dominant herbivores remained the same for all sites across the gradient, but associations of characteristic herbivores changed between high and low pollution sites. Additionally, five decades of air pollution research in the San Bernardino Mountains have confirmed many air pollution impacts already affecting soil processes and plant growth (Miller and McBride, 1999; Bytnerowicz and Fenn, 1996; Fenn et al., 1996). Oak herbivores showed strong differences along the pollution gradient with communities at high pollution sites dominated by chewing insects but there were no significant differences among sucking insects along the gradient. To conclude, while diversity and richness have traditionally been used to monitor environmental impacts on ecological communities, had we relied on these measures, we would have detected no effect of air pollution on herbivore communities along the air pollution gradient in the San Bernardino Mountains. In spite of uniform herbivore richness and diversity across the gradient, there were shifts in herbivore community organization that were strongly associated with air pollution exposure, and that have the potential to affect forest processes. These shifts in herbivore communities could affect resources for higher trophic levels as well as influence rates of nutrient cycling. Monitoring changes in oak herbivore communities through discriminant function analysis may serve as an important indicator for assessing the effects of air pollution in xeric western forests.
Acknowledgements We are grateful to Stephanie Washburn, Chris Hanlon, Stuart Wooley, Sarah Allen, Idara Essien, and Jocelyn Holt for their assistance in the field and laboratory. We thank J. Daniel Hare, Richard Redak, Edith Allen, Rocio Alonso and Mark Fenn for their comments on the manuscript. Funding was provided in part by the University of California Graduate Division through a dissertation improvement grant and by a grant from the University of California Academic Senate.
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