Nitrogen deposition in Mediterranean forests

Nitrogen deposition in Mediterranean forests

Environmental Pollution 118 (2002) 205–213 www.elsevier.com/locate/envpol Nitrogen deposition in Mediterranean forests F. Roda`*, A. Avila, A. Rodrig...

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Environmental Pollution 118 (2002) 205–213 www.elsevier.com/locate/envpol

Nitrogen deposition in Mediterranean forests F. Roda`*, A. Avila, A. Rodrigo Center for Ecological Research and Forestry Applications (CREAF), Universitat Auto`noma de Barcelona, E-08193 Bellaterra, Spain Received 10 January 2001; accepted 31 August 2001

‘‘Capsule’’: Measurements of total atmospheric deposition indicate the retention of N within broadleaved evergreen forests. Abstract Atmospheric deposition of inorganic nitrogen was studied at two forested sites in the Montseny mountains (northeast Spain), peripheral to the Barcelona conurbation, and at a nearby lowland town, using bulk deposition, wet-only deposition, throughfall, and dry deposition inferred from branch-washes and surrogate surfaces (metacrylate plates). Bulk deposition inputs of ammonium and nitrate did not show significant temporal trends over a 16-year period. Bulk inputs of inorganic N were moderate, ranging from 6 to 10 kg N ha1 year1 depending on the time period considered and the degree of site exposure to polluted air masses from the Barcelona conurbation. Large dry-sedimented particles played a minor role, since wet-only inputs were virtually identical to bulk inputs. On the contrary, branch- and plate-washes indicated substantial dry inputs of N gases and small particles. Total atmospheric deposition was estimated at 15–22 kg N ha1 year1, most of it being retained within the studied broadleaved evergreen forests. Ecosystem N availability is thus likely to be increasing in these forests. # 2002 Elsevier Science Ltd. All rights reserved.

1. Introduction Atmospheric deposition of nitrogen to forest ecosystems is rising in recent years due to the increased emissions of oxidised N from industrial and transport activities, and of reduced N from intensive livestock production and agriculture (Galloway, 1995; Vitousek et al., 1997). Increased deposition of atmospheric N compared to pre-industrial levels has been documented for northern Europe (Brimblecombe and Pitman, 1980; Sutton et al., 1993) and for some sites in the United States (Fenn et al., 1998). In the eastern United States, the release of NOx from fossil fuel consumption has increased the atmospheric deposition of N by roughly 10-fold compared to pre-industrial times (Hicks et al., 1990). Future emissions of N compounds are expected to raise worldwide because of the enhanced use of nitrogenous fertilisers and fossil fuel consumption, and this increase is expected to be particularly intense in Asia and Africa (Galloway, 1995). N deposition threatens

* Corresponding author. Tel.: +34-93-581-1928; fax: +34-93-5811312. E-mail address: [email protected] (F. Roda`).

biodiversity in nutrient-poor ecosystems (Bobbink et al., 1998) and contributes to soil and freshwater acidification in acid-sensitive areas and to freshwater and marine eutrophication (Paerl, 1995; Vitousek et al., 1997). Enhanced N deposition to forests has exceeded the assimilation capacity of some ecosystems leading to forest nutrient imbalances and NO 3 leaching to streamwaters, in a process known as ‘‘nitrogen saturation’’ (Aber et al., 1989, 1998; Agren and Bosatta, 1988; Nihlgard, 1985). Nitrogen saturation has been described for forests receiving large loads of N, ranging between 20 and 100 kg N ha1 year1, in the Netherlands (Van Breemen et al., 1987; Wyears et al., 1992), Great Britain (Emmet et al., 1993), Scandinavia (Andersen et al., 1993) and the United States (Aber et al., 1989, 1998; Bytnerowicz and Fenn, 1996; Fenn et al., 1996, 1998). In Catalonia (northeastern Spain) increasing N emissions are also occurring from increased fossil fuel consumption and livestock production (Anon., 1991). However, little is known on the magnitude and trends of atmospheric N deposition for rural areas in the Mediterranean Basin, and on the potential impact of N deposition on Mediterranean forests. In this paper, we use our results obtained in several studies conducted since 1983 in the Montseny mountains (Catalonia) to:

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 (1) describe the current NH+ 4 and NO3 concentrations in rainwater and their bulk deposition rates; (2) give an estimate of dry N deposition for two evergreen broadleaved Mediterranean forests, dominated by holm oak (Quercus ilex L.), differentially exposed to polluted air masses; and (3) discuss the effects of atmospheric inputs on the N cycle in these holm oak forests.

2. Site description and methods The Montseny mountains (41 460 N, 2 210 E; 40 km NNE from Barcelona) cover an area of about 400 km2, reach their highest altitude at 1707 m a.s.l., and are mostly forested (holm oak, pines, and beech), with heathlands and grasslands in the upper parts and arable crops in the lower reaches. The bedrock consists mostly in metamorphic phyllites and schists. The climate is montane Mediterranean. Mean annual precipitation at our study sites ranges from 770 to 900 mm, falling mostly as rainfall. Mean annual temperature ranges from 10 to 14  C. We have studied atmospheric deposition at three study sites at Montseny along an altitudinal and pollution gradient (Fig. 1), the latter originating from differential exposure to the polluted atmosphere of the urban and industrial area around Barcelona. One sampling site, named LC was at La Castanya valley (altitude 700 m a.s.l.), central to the Montseny massif, in a topographically sheltered position from polluted air masses from the Barcelona conurbation, and removed from local pollution sources. A second forested site, named RP (Riera de Sant Pere, altitude 535 m) was located in the southern reaches of Montseny in a more exposed position to polluted air from the Barcelona conurbation. A third site, named PA (Santa Maria de Palautordera, altitude 200 m) was located in the lowlands at the foot of the mountains, in the outskirts of a town of 5000 people, in a mosaic landscape of cultivated fields, livestock farms, forests, roads and industries (Fig. 1).

The two forested sites (LC and RP) were located amidst extensive holm oak forests, bulk deposition being sampled in clearings of sufficient diameter to avoid interference by trees. The main topographic characteristics and the stand structure of these forests are summarised in Table 1. Both holm oak forests were structurally matched as closely as feasible. Linear distance between PA and RP is 5.3 km, while RP and LC lie 7.2 km apart. The lowland site (PA) lies 3.5 km away from a major motorway, and 15–17 km from two mid-size cities (50 000–100 000 people) which lie to the southwest and south. See breezes blowing from these directions bring polluted air to PA and RP sites. Furthermore, the lowland site is often affected by thermal inversions that reduce pollutant dispersal. Bulk precipitation was collected weekly at each site with two replicate collectors consisting of a funnel connected to a 10-l polyethylene bottle. At PA, wet-only deposition was also sampled, using an automatically opening-and-closing collector. Results of the comparison of bulk and wet-only deposition for 4 years (1995–1999) for this site indicated very small and nonsignificant differences in ion concentrations between Table 1 Site characteristics (meanS.E., n=4 per site) of the two studied holm oak forests at Montseny (LC and RP)a Variable

Altitude (m) Slope ( ) Stand density (stems ha1) Basal area (m2 ha1) Tree height (m) Canopy depth (m) Tree cover (%)

Forest study site LC

RP

7314.3 37.50.6 212716.2 26.42.0 5.90.3 3.10.2 1280.2

535 2.9 20.5 0.9 1754103 22.3 0.5 6.4 0.3 4.0 0.3 96 0.1

P

<0.0001 <0.0001 0.01 ns ns 0.04 ns

a Stand and tree data are for stems having a diameter at breast height >5 cm. Differences between sites were tested through a paired Student-t test of two tails.

Fig. 1. Idealised topographical profile showing the three sampling locations in the Montseny mountains. Exposure to polluted air masses and to local pollution sources decreases from the lowland site (Palautordera) through the low-elevation forest site (RP, Riera de Sant Pere) to the midelevation and topographically sheltered forest site (LC, La Castanya). For clarity, a deep valley between RP and LC has been omitted. See the text for relevant distances.

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both methods of precipitation collection, the only significant difference being for NH+ 4 that was 5% higher in wet-only precipitation (Avila and Rodrigo, 2001). Also, the analysis of the dry deposition recovered from washing the funnels during rainless weeks indicated that the flux of dry-sedimented particles onto the funnels was small, generally below 10% of bulk deposition amounts (Avila, 1996; Rodrigo, 1998). Therefore, we will use here the bulk deposition measurements as representative of rain chemistry and deposition. Throughfall was weekly sampled in four replicate plots inside the holm oak forests at LC and RP. The plots were circular, with 7-m radius. Each plot contained eight throughfall collectors, consisting of a 10-cm diameter funnel connected to a 2-l polyethylene bottle. The sampling bottles, both for throughfall and bulk deposition, were kept in the dark to minimise biological transformations. Stemflow was also sampled as described in Rodrigo (1998). As stemflow N fluxes were very minor compared to throughfall these results are not commented on here, but where relevant throughfall+stemflow fluxes are given. An experiment of branch- and plate-washings was undertaken at LC and RP to collect substances drydeposited onto the canopy during rainless periods. In each site, eight extruding branches at the top of dominant trees were chosen for the experiment and marked. On five rain-free periods they were washed in situ with distilled deionised water after exposure times ranging between 72 and 187 h since the last rainfall. After the fifth period, they were cut and carried to the laboratory to measure the supported leaf area. For the same exposure periods, six metacrylate plates of 2020 cm and 0.2 mm thickness were horizontally exposed on top of adjacent trees. The material deposited over the exposed surface was extracted in the laboratory with a 10-min deionised water wash. More details can be found in Rodrigo and Avila (2001). Study periods differed among sites and among the studied processes. The longest record was for rain chemistry at LC, where 16 years of data were available (August 1983–August 1999). For PA, the sampling period was from June 1995 to June 1999, and for RP from 6 June 1995 to 25 June 1996. Throughfall used here from LC and RP are also for the latter period. The washing experiment of branches and plates was conducted in May–June 1996. All water samples (bulk deposition, wet deposition, throughfall, and leaf and plate washes) were taken on the day of sampling to the laboratory where pH, conductivity and alkalinity were analysed within 24–48 h. Samples were then filtered with 0.45-mm Millipore filters, and the filtrate was deep-frozen for later analyses   of major anions (SO2 4 , NO3 and Cl ) and cations + + + 2+ 2+ (Na , K , Ca , Mg , NH4 ). Nitrate was analysed by ion chromatography and NH+ 4 by flow injection

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analysis until 1996 and by ion chromatography thereafter. All ion analyses included synthetic samples of known concentrations within the runs to check for precision and accuracy. Data were further screened for analytical quality by the ratios sum of cations/sum of anions (equivalent basis) and calculated conductivity/ measured conductivity.

3. Results and discussion + 3.1. Concentrations of NO 3 and NH4 in precipitation

The rain chemistry at Montseny was characterised by a positive alkalinity with most of the potential acidity of strong acids being neutralised by Ca2+ and NH+ 4 (Roda` et al., 1993; Avila, 1996; Rodrigo, 1998). In contrast with the rain chemistry in central and northern European sites, where NH+ 4 is the main base counterpart to acidic anions, here Ca2+ had a prominent role. This is the outcome of calcium carbonate being incorporated into the atmosphere from edaphic, urban or industrial sources in Spain, and from the transport of African dust in sporadic but strongly alkaline red rains (Avila et al., 1997; Avila and Alarco´n, 1999). Annual mean concentrations of inorganic N in rainwater at the two Montseny forested sites were moderate (Table 2), those  of NH+ 4 being slightly higher than NO3 . Higher mean + NH4 concentrations were found at the lowland site (PA), which was surrounded by arable land and livestock farms. For the year of joint sampling, mean nitrogen concentrations were 17–24% higher in the exposed forested site compared to the sheltered one (a non-significant difference, paired two-tailed t-test), but as rainfall varied in the opposite direction bulk deposition fluxes were equal at both sites (Table 2). 3.1.1. Time trends At Montseny, the participation of SO2 as an acid 4 anion in rainwater was much higher than that of NO 3, but this difference showed a decreasing trend due to the decline of SO2 4 in precipitation and the maintenance of NO 3 concentrations (Fig. 2). Bulk deposition data for the 16-year period at LC showed no significant temporal trends in the concentrations of either NO 3 or NH+ 4 (Fig. 3). Average rainwater concentrations at one particular site are a combination of the contributions of longrange transport with the medium-range and local contributions. In a study of the chemical features of precipitation at Montseny considering the meteorological provenances of air masses, Avila and Alarco´n (1999) classified the precipitation in four categories: oceanic, local, European and African. Rains coming from the European continent presented the highest  + concentrations for H+, excess-SO2 4 , NO3 , and NH4 .

F. Roda` et al. / Environmental Pollution 118 (2002) 205–213

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Table 2 Volume-weighted mean concentrations in bulk or wet-only precipitation and deposition of inorganic N at Montseny Site

Length of study period (year)

Precipitation (mm year1)

Concentration (meq l1)

Deposition (kg N ha1 year1)

NH+ 4

NO 3

NH4-N

NO3-N

LC, bulk

16a

929

22.8

20.9

2.96

2.71

RP, bulk LC, bulk

1b 1b

1048 1275

35.2 30.0

32.7 26.4

5.17 5.36

4.80 4.71

PA Bulk Wet

4c 4c

740 705

44.7 47.0

32.4 31.7

4.63 4.64

3.36 3.13

a b c

From August 1983 to August 1999. From 6 June 1995 to 25 June 1996. From June 1995 to June 1999.

Fig. 2. Temporal decrease in the ratio between mean annual volumeweighted concentrations of sulphate and nitrate in bulk deposition at Montseny (La Castanya site) during a 16-year study period. Each dot corresponds to one hydrological year, starting on 1 August.

Nitrate and NH+ 4 concentrations in European events were roughly double than those of other provenances (Table 3). However, events of European origin declined in magnitude during the study period, and as a consequence, a significant decrease in N deposition from European events was found for the period 1983–1994 (Table 4). 3.2. Dry deposition of N to holm oak forests Forest canopies act as effective deposition surfaces for airborne substances (e.g. Parker, 1983; Lindberg et al., 1986; Lovett, 1994), and throughfall and stemflow fluxes have been often used to describe the total atmospheric inputs of some elements (particularly S), assuming that they include wet and dry deposition inputs from the atmosphere (Mayer and Ulrich, 1972; Parker, 1983). However, the chemistry of throughfall and stemflow can also be modified by internally leached ions or through direct uptake at plant surfaces. In the latter case, if input fluxes (dry and wet) are lower than canopy

Fig. 3. Annual mean volume-weighted concentrations of (a) NH+ 4 and (b) NO 3 at La Castanya (Montseny) during a 16-year study period. Neither ion showed a significant temporal trend. Each dot corresponds to one hydrological year, starting on 1 August.

uptake, the net fluxes in throughfall and stemflow will be negative, implying a net retention in the canopy. This is often the case for N fluxes in sites with moderate N inputs (Bellot and Escarre´, 1991; Lovett and Lindberg, 1993; Pucket, 1990; Schlesinger et al., 1982), and precludes the use of throughfall and stemflow for the estimation of dry deposition fluxes of N. At Montseny, N inputs in rainwater were moderate + (Table 2). At the sheltered site (LC), NO 3 and NH4 concentrations decreased as rainwater crossed the

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Table 3 Volume-weighted mean ion concentrations for four meteorological classes of rainfall provenance at La Castanya (LC) Montseny, 1983–1994a Rainfall provenance

% Annual precipation

Conductivity

Alkalinity

Na+

K+

Ca2+

Mg2+

NH+ 4

NO 3

SO2 4

Cl

Oceanic Local European African

60.1 15.8 4.5 19.6

13.7 18.5 35.5 26.7

14.8 1.7 41.0 82.7

15.4 20.7 32.1 42.4

2.7 4.4 5.4 7.3

23.4 53.1 70.9 162

6.2 10.5 16.6 18.7

20.1 30.5 57.2 17.7

16.8 26.7 49.1 21.4

35.3 57.7 103 56.8

20.0 26.3 38.5 52.8

a Percentage of annual precipitation contributed for each provenance is given. Units are mS cm1 (at 20  C) for conductivity, and meq L1 for ion concentrations. Adapted from Avila and Alarco´n (1999).

Table 4 Mean annual bulk deposition of N (kg N ha1 year1) at La Castanya (LC, Montseny), classified by rainfall provenance and time perioda Time period

Oceanic

Local

European

African

First 5-year period Second 5-year period P-value

2.68 2.76 0.87

1.01 1.27 0.33

0.86 0.39 0.02

0.91 0.88 0.94

a The first period comprises the first 5 study years (August 1983–August 1988), and the second period comprises the next five years (August 1989– August 1994). P-values testing for differences in N inputs between both periods are also shown.

canopy, indicating a net retention in the canopy. However, at the forested site more exposed to polluted air + masses (RP) NO 3 and NH4 concentrations increased in throughfall, compared to bulk precipitation, more than explained by the interception losses of water, indicating that dry deposition exceeded canopy uptake at this site. Throughfall concentrations and fluxes of both NO 3 and NH+ 4 were significantly higher in the exposed than in the sheltered site, suggesting higher dry deposition of N at the former. In terms of fluxes, negative net fluxes for + NO 3 and NH4 were obtained at LC and positive, though small fluxes at RP (Fig. 4). However, since N can be taken up in the canopy by microbes, epiphytes, and tree leaves, the extent of the dry deposition flux cannot be estimated from net throughfall fluxes, but a comparison can be made between sites to see the relative role of dry deposition. Based on the element balance equation for the canopy:

For NO 3: Dlc ¼ UC þ ð1:64Þ  4:71 ¼ CU  6:35 At the exposed site: DDrp ¼ CU þ NTrp  Wrp For NH+ 4 : Drp ¼ CU þ 1:16  5:17 ¼ UC  4:01 For NO 3: Drp CU þ 0:25  4:80 ¼ UC  4:55 If both sites have similar canopy uptake, then For NH+ 4 :

WD þ DD ¼ CU þ NT

Drp ¼ Dlc þ 2:97

Where WD and DD are wet deposition and dry deposition element fluxes (kg ha1 year1), respectively, CU is canopy uptake and NT is net throughfall. Solving for dry deposition, we can write the equation at each site and for each inorganic N species: At the sheltered site (figures are for the year of joint sampling, in kg N ha1 year1):

and for NO 3:

DDlc ¼ CU þ NTlc  Wlc For NH+ 4 : Dlc ¼ CU þ ð1:61Þ  5:36 ¼ CU  6:97

Drp ¼ Dlc þ 1:8 Assuming similar canopy uptake fluxes at both sites (a reasonable assumption given the similarities in soil type and stand structure), the exposed site would therefore receive a dry deposition flux greater by 3 and 2 kg  N ha1 year1 for NH+ 4 and NO3 , respectively, than the sheltered site. To estimate dry deposition fluxes at Montseny, we used another approach based on a technique of washing branches and surrogate surfaces (Rodrigo and Avila,

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Fig. 4. N fluxes in bulk deposition, throughfall (plus stemflow), and net throughfall in a topographically sheltered (La Castanya, LC) and an exposed (Riera de Sant Pere, RP) site at Montseny. (a)  NH+ 4 and (b) NO3 . Net throughfall is taken here as the difference between bulk deposition and throughfall+stemflow. Data are for the period 6 June 1995 to 25 June 1996.

2001). For five sampling occasions on May–June 1996, dry deposition fluxes per unit leaf or plate area (mg cm2 h1) for NO3-N and NH4-N were obtained at each site, based on the quantities recovered by washing treetop branches and metacrylate plates (Table 5). These values were extrapolated to estimate the dry deposition on the canopy at an annual scale (Table 6), assuming a leaf area index (LAI) of five for both holm oak sites (Sabate´ et al., 1999). This extrapolation has to be borne in mind since there are some sources of uncertainty in the deposition fluxes linked to: (1) NH3 emissions being greater than average during the experiment period because of fertiliser application in late winter and spring, and (2) a probable gradient of dry deposition decrease within the canopy.

In the absence of more data, we will take these values as a preliminary estimate of dry deposition fluxes. Deposition was higher on branches than on plates for NO3-N and NH4-N at both sites (Table 5). The recovery of NH+ 4 from plates was below our detection limit (< 2.5 meq l1). Leaf-bearing branches probably were thus much more efficient than plates in capturing airborne substances, the difference being lower for large particles (and hence for NO 3 ) than for gases and small particles (and hence for NH+ 4 ). The much smaller size of holm oak leaves (3–6 cm2) compared to plates (400 cm2) and their higher surface roughness contributed to this result. Therefore, and since inorganic N is not easily leached from healthy plants, we believe that branch washes give a better estimate of dry-deposited N than metacrylate plates do. Nonetheless, branch-washes may overestimate NH+ 4 deposition if labile organic N compounds leached from branch surfaces release NH+ 4 to the washing solution. Dry deposition fluxes estimated from the washing of branches were higher than bulk deposition fluxes at both sites (Table 6). Total deposition of inorganic N was 14.8 and 21.9 kg ha1 year1 at the sheltered (LC) and exposed (RP) sites, respectively. With these data, dry deposition accounts for 62–67% of total N deposition, even without taking into account dry deposition processes that are underrepresented in our data, such as stomatal uptake of N gases. This percentage is similar to the dry deposition contribution reported for a drier EMEP (European Monitoring and Evaluation Programme) site in Tortosa (southernmost Catalonia) where dry deposition contributed 60% of total N deposition (Ferna´ndez-Patier et al., 1990). However, these authors estimated dry deposition through a different method, combining measurements of air con+ centrations of NO2 and NO 3 and NH4 particles with literature values of deposition velocities. For the low elevation forests within the Integrated Forest Study in eastern United States (Johnson and Lindberg, 1992), dry deposition of N compounds (HNO3, and NO 3 and NH+ 4 in fine and coarse particles) ranged between 39 and 59% of total deposition (Lovett and Lindberg, 1993). Unfortunately, atmospheric concentrations of Ncontaining gases and particles were not available at our sites, or at comparable rural sites in Catalonia. 3.3. The nitrogen cycle in holm oak forests Nutrient cycling has been intensively studied in the holm oak forest at LC (Roda` et al., 1999a). Soil nitrogen availability is moderately high, with 80 kg N ha1 year1 being mineralised in the first 20 cm of soil (Bonilla and Roda`, 1992). However, due to the low N concentrations in the sclerophyllous leaves of holm oak, and to foliar retranslocation before leaf abscission the N flux in litterfall is relatively low (37 kg N ha1 year1,

F. Roda` et al. / Environmental Pollution 118 (2002) 205–213 Table 5 Mean dry deposition rates on treetop branches and metacrylate plates obtained from five washing experiments after rainless periodsa Site

Sheltered (LC) Exposed (RP) a

Nitrate

Ammonium

Branches

Plates

Branches

Plates

11.5 14.8

6.3 8.4

9.4 18.5

ind ind

2

211

Table 6 Bulk and dry N deposition at two forested sites at Montsenya Dry

Totalb

Site

Bulk

Sheltered (LC) NH4-N NO3-N

2.96 2.71

4.12 5.04

Total

5.67

9.16

14.8

62

Exposed (RP) NH4-N NO3-N

3.81c 3.54c

8.08 6.48

11.9c 10.0c

68 65

Total

7.35c

21.9c

67

7.08 7.75

% Dry/total 58 65

1

Figures are mg N cm h , expressed on a leaf area basis for branches and on plate area for plates.

Escarre´ et al., 1999). Nitrogen cycles tightly within the holm oak forest, with dissolved N losses in streamwater being < 0.1 kg N ha1 year1 (Avila et al., 1999) and particulate losses being also low in these thickly forested catchments. Therefore, as is usually the case for undisturbed forests, the within-ecosystems N fluxes are much larger than either N input or outputs. Biological demand for N is a major determinant in this closed N cycle, as clearly demonstrated by a roottrenching experiment, where water and nutrient uptake by tree roots was prevented: NO 3 concentration in soil water, which was very low in control plots, skyrocketed in trenched plots to 2.5 times above European drinking water standards (Bonilla, 1990; Bonilla and Roda`, 1990). The highly regulated cycle of N in holm oak forests is also revealed by the ecosystem solution profile: NO 3 concentrations in mineral–soil water and streamwater are much lower than in rainwater and throughfall (Serrasolses et al., 1999). Peaks of NO 3 export in streamwater only occurred shortly during stormflows (Avila et al., 1992). In contrast to cold-temperate forests, streamwater outputs of NO 3 remained low even in autumn and winter in holm oak forests (Pin˜ol et al., 1992), probably because biological activity can proceed all year round in these forests. 3.4. Significance of N deposition for the N cycle in Mediterranean holm oak forests Our results indicate that at Montseny, the total atmospheric deposition of inorganic N on holm oak forests is not negligible (between 15 and 22 kg N ha1 year1), though it is much lower than in northwestern Europe, eastern North America, and California. Nevertheless, net throughfall fluxes were negative at the sheltered site and positive but very small at the exposed site. Given the substantial dry deposition flux, this reveals that canopy uptake of atmospherically deposited N takes places at both forests, suggesting that they have not yet reached N saturation. Inclusion of organic N (not analysed in this study) would increase substantially our estimate of N deposition, since organic N often contributes 20–30% of total N in rainwater (Peierls and Paerl, 1997; Scudlark et al., 1998).

14.6

a Bulk deposition data are for the period August 1983–August 1999 at the sheltered site (LC) and for the period 6 June 1995–25 June 1996 at the exposed one (RP). Dry deposition data are extrapolated from five branch-washing periods in May–June 1996 (see text). Units are kg N ha1 year1, except for the last column (%). b Total=Bulk+dry. Dry deposition on branches is added to bulk deposition because the latter corresponds mostly to wet deposition (see text). c Rainfall was extremely high during the one-year study period at RP (1048 mm at RP; 1275 mm at LC, which is 37% higher than the 16-year average at LC). Therefore, bulk deposition fluxes for RP in this table have been corrected to an average-rainfall year multiplying the volume-weighted mean concentrations at RP by the estimated mean annual rainfall at RP (772 mm). The lack of relationship between annual rainfall and the mean annual weighted concentrations  of NH+ 4 and NO3 in the 16-year record at LC (r=0.39 and r=0.20, respectively) supports this procedure.

In a comparative study of input output N budgets for 65 forested catchments and plots across Europe, Dise and Wright (1995) proposed a N deposition threshold of about 10 kg N ha1 year1 in throughfall fluxes upon which significant NO 3 leaching would start to occur from forests. With throughfall exceeding 25 kg N ha1 year1, the ecosystems would become N saturated and would start to leach large quantities of nitrogen. Throughfall plus stemflow fluxes of N at our sites were around the former threshold: 6.8 kg N ha1 year1 at the sheltered site, and 11.4 at the exposed one (Fig. 4). Recent studies in California also provided evidence for N saturation with N deposition rates around 25 kg N ha1 year1 (Bytnerowicz and Fenn, 1996; Fenn et al., 1998). The importance of atmospheric deposition in the N dynamics of holm oak forests can be evaluated by comparing the magnitude of the input flux with the intrasystemic nutrient fluxes. Here we will use the annual N retention in new aboveground woody tissues as the relevant flux for indicating nutrient accumulation in new woody biomass. This flux has been estimated at 4.6 kg N ha1 year1 for the LC forest (Escarre´ et al., 1999). Therefore, with total atmospheric inputs of inorganic N ranging between 15–22 kg N ha1 year1, atmospheric deposition would account for 3–5 times

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the N yearly incorporated in the aboveground woody biomass. Expressed in terms of the N uptake needed for sustaining current aboveground production (41.6 kg N ha1 year1, Escarre´ et al., 1999), atmospheric deposition supplies a N flux equivalent to 36–53% of this amount. Holm oak forests in the Mediterranean Basin have been subjected for centuries to repeated nutrient losses through fire and biomass harvesting. These disturbances may have depleted the soil nutrient pools. Nutrient limitation may be partly responsible for the low net aboveground primary production of holm oak forests (6.3 t ha1 year1 at LC; Iba`n˜ez et al., 1999), although water limitation and other factors are also involved. An experiment conducted in the drier holm oak forest of Prades (120 km to the southwest of Montseny) revealed that net aboveground production of trees increased both after irrigation and fertilisation with 250 kg N ha1 (Roda` et al., 1999b). N fertilisation greatly increased leaf and acorn production, demonstrating that Mediterranean forests can respond to increased N availability despite their strong water limitation. Of course, the effect of continued additions of N at low doses, such as atmospheric deposition at Montseny, may have completely different effects from those of a single, large pulse of N fertiliser. However, since the weight of evidence indicates that atmospherically deposited N is being retained within the ecosystem, a build-up of N pools in the soil and vegetation is likely. Relevant enough, the holm oak forests of LC and RP are now accreting biomass in a recovering phase after the last forest harvest. Atmospheric N inputs may be used by trees to sustain this growth, but the long-term effects of the enhanced N availability on the forest ecosystem are a matter of concern.

References Aber, J.D., McDowell, W., Nadelhoffer, K., Magill, A., Berntson, G., Kamakea, M., McNulty, S., Currie, W., Rustad, L., Fernandez, I., 1998. Nitrogen saturation in temperate forest ecosystems. BioScience 48, 921–934. Aber, J.D., Nadelhoffer, K.J., Steudler, P., Melillo, J.M., 1989. Nitrogen saturation in northern forest ecosystems. BioScience 39, 378–386. Agren, G.I., Bosatta, E., 1988. Nitrogen saturation of terrestrial ecosystems. Environmental Pollution 54, 185–198. Andersen, H.V., Hovmand, M.F., Hummelshoj, P., Jensen, N.O., 1993. Measurement of ammonia flux to a spruce stand in Denmark. Atmospheric Environment 27, 189–202. Anonymous, 1991. La Contaminacio´ Atmosfe`rica. Factors Meteorolo`gics, Contaminants i Efectes de la Contaminacio´ atmosfe`rica. Generalitat de Catalunya, Departament de Polı´tica Territorial i Obres Pu´bliques, Barcelona. Avila, A., 1996. Time trends in the precipitation chemistry at a montane site in northeastern Spain for the period 1983–1994. Atmospheric Environment 30, 1363–1373.

Avila, A., Alarco´n, M., 1999. Relationship between precipitation chemistry and meteorological situations at a rural site in NE Spain. Atmospheric Environment 33, 1663–1677. Avila, A., Bellot, J., Pin˜ol, J., 1999. Element budgets in catchments. In: Roda`, F., Retana, J., Gracia, C., Bellot, J. (Eds.), Ecology of Mediterranean Evergreen Oak Forests. Springer, Berlin, pp. 283–296. Avila, A., Pin˜ol, J., Roda`, F., Neal, C., 1992. Storm solute behaviour in a montane Mediterranean forested catchment. Journal of Hydrology 140, 143–161. Avila, A., Queralt-Mitjans, I., Alarco´n, M., 1997. Mineralogical composition of African dust delivered by red rains over northeastern Spain. Journal of Geophysical Research 102 (D18), 21977–21996. Avila, A., Rodrigo, A., 2001. Bulk and wet-only chemistry at a rural site in northeastern Spain. Journal of Hydrological Science. (submitted for publication). Bellot, J., Escarre´, A., 1991. Chemical characteristics and temporal variations of nutrients in throughfall and stemflow of three species in a Mediterranean holm oak forest. Forest Ecology and Management 41, 125–135. Bobbink, R., Hornung, M., Roelofs, J.G.M., 1998. The effects of airborne nitrogen pollutants on species diversity in natural and seminatural European vegetation. Journal of Ecology 86, 717–738. Bonilla, D., 1990. Biogeoquı´mica del suelo en un encinar y landas del Montseny: la solucio´n del suelo, la dina´mica del nitro´geno y la respuesta a una perturbacio´n experimental. PhD dissertation, Autonomous University of Barcelona, Bellaterra. Bonilla, D., Roda`, F., 1990. Nitrogen cycling responses to disturbance: trenching experiments in an evergreen oak forest. In: Harrison, A.F., Ineson, P., Heal, O.W. (Eds.), Nutrient Cycling in Terrestrial Ecosystems. Elsevier, London, pp. 179–189. Bonilla, D., Roda`, F., 1992. Soil nitrogen dynamics in a holm oak forest. Vegetatio 99/100, 247–257. Brimblecombe, P., Pitman, J., 1980. Long-term deposit at Rothamsted, Southern England. Tellus 32, 261–267. Bytnerowicz, A., Fenn, M.E., 1996. Nitrogen deposition in California forests: a review. Environmental Pollution 92, 127–146. Dise, N.B., Wright, R.F., 1995. Nitrogen leaching from European forests in relation to nitrogen deposition. Forest Ecology and Management 71, 153–161. Emmet, B.A., Reynolds, B., Stevens, P.A., Norris, D.A., Hughes, S., Go¨rres, J., Lubrecht, I., 1993. Nitrate leaching from afforested Welsh catchment: interaction between stand age and nitrogen deposition. Ambio 22, 386–394. Escarre´, A., Carratala´, A., Avila, A., Bellot, J., Pin˜ol, J., Milla´n, M., 1999. Precipitation chemistry and air pollution. In: Roda`, F., Retana, J., Gracia, C., Bellot, J. (Eds.), Ecology of Mediterranean Evergreen Oak Forests. Springer, Berlin, pp. 195–208. Fenn, M.E., Poth, M.A., Aber, J.D., Baron, J.S., Bormann, B.T., Johnson, D.W., Lemly, A.D., McNulty, S.G., Ryan, D.F., Stottlemyer, R., 1998. Nitrogen excess in North American ecosystems: predisposing factors, ecosystem responses, and management strategies. Ecological Applications 8, 706–733. Fenn, M.E., Poth, M.A., Johnson, D.W., 1996. Evidence for nitrogen saturation in the San Bernardino mountains in southern California. Forest Ecology and Management 82, 211–230. Ferna´ndez-Patier, R., de la Serna, J., Esteban Lefler, M., Dı´ez Herna´ndez, P., Garcı´a Sa´nchez, J., 1990. Dry and wet deposition of nitrogen containing species. In: Beilke, S., Milla´n, M., Angeletti, G. (Eds.), Field Measurement and Interpretation of Species Derived from NOx, NH3 and VOC Emissions in Europe (Air Pollution Research Report 25). Commission of the European Community. Galloway, J.N., 1995. Acid deposition: perspectives in time and space. Water, Air, and Soil Pollution 85, 15. 24. Hicks, B.B., Draxler, R.R., Albritton, D.L., Fehensfeld, F.C., Dodge, M., Schwartx, S.E., Tanner, R.L., Hales, J.M., Meyers, T.P., Vong, R.J., 1990. Atmospheric Processes Research and Process Model Development. Acidic Deposition. State of Science and Technology

F. Roda` et al. / Environmental Pollution 118 (2002) 205–213 (Report 2). National Acid Precipitation Assessment Program, Washington, DC. Iba`n˜ez, J.J., Lledo´, M.J., Sa´nchez, J.R., Roda`, F., 1999. Stand structure, aboveground biomass and production. In: Roda`, F., Retana, J., Gracia, C., Bellot, J. (Eds.), Ecology of Mediterranean Evergreen Oak Forests. Springer, Berlin, pp. 31–45. Johnson, D.W., Lindberg, S.E. (Eds.), 1992. Atmospheric Deposition and Forest Nutrient Cycling. Springer, New York. Lindberg, S.E., Lovett, G.M., Richter, D.D., Johnson, D.W., 1986. Atmospheric deposition and canopy interactions of major ions in a forest. Science 231, 141–145. Lovett, G.M., 1994. Atmospheric deposition of nutrients and pollutants in North America: an ecological perspective. Ecological Applications 4, 629–650. Lovett, G.M., Lindberg, S.E., 1993. Atmospheric deposition and canopy interactions of nitrogen in forests. Canadian Journal of Forest Research 23, 1603–1616. Mayer, R., Ulrich, B., 1972. Conclusions on the filtering action of forests from ecosystem analysis. Oecologia Plantarum 9, 157–168. Nihlgard, B., 1985. The ammonium hypothesis—an additional explanation for the forest dieback in Europe. Ambio 14, 2–8. Paerl, H.W., 1995. Coastal eutrophication in relation to atmospheric nitrogen deposition: current perspectives. Ophelia 41, 237–259. Parker, G.G., 1983. Throughfall and stemflow in the forest nutrient cycle. Advances in Ecological Research 13, 57–133. Peierls, B.L., Paerl, H.W., 1997. Bioavailability of atmospheric organic nitrogen deposition to coastal phytoplankton. Limnology and Oceanography 42, 1819–1823. Pin˜ol, J., Avila, A.F., Roda`, 1992. The seasonal variation of streamwater chemistry in three forested Mediterranean catchments. Journal of Hydrology 140, 119–141. Pucket, L.J., 1990. Estimates of ions sources in deciduous and coniferous throughfall. Atmospheric Environment 24, 545–555. Roda`, F., Bellot, J., Avila, A., Escarre´, A., Pin˜ol, J., J. Terradas., 1993. Saharan dust and the atmospheric inputs of elements and alkalinity to mediterranean ecosystems. Water, Air, and Soil Pollution 66, 277-288. Roda`, F., Retana, J., Gracia, C., Bellot, J. (Eds.), 1999a. Ecology of Mediterranean Evergreen Oak Forests. Springer, Berlin. Roda`, F., Mayor, X., Sabate´, S., Diego, V., 1999b. Water and nutrient limitations to primary production. In: Roda`, F., Retana, J., Gracia,

213

C., Bellot, J. (Eds.), Ecology of Mediterranean Evergreen Oak Forests. Springer, Berlin, pp. 183–194. Rodrigo, A., 1998. Deposicio´ atmosfe`rica en dos alzinars (Quercus ilex L.) del Montseny sotmesos a una exposicio´ contrastada als contaminants de l’a`rea barcelonina i vallesana. PhD dissertation, Autonomous University of Barcelona, Bellaterra. Rodrigo, A., Avila, A., 2001. Dry deposition to the forest canopy and surrogate surfaces in two Mediterranean holm oak forests in Montseny (Spain). Water, Air, and Soil Pollution (submitted for publication). Sabate´, S., Sala, A., Gracia, C., 1999. Leaf traits and canopy organization. In: Roda`, F., Retana, J., Gracia, C., Bellot, J. (Eds.), Ecology of Mediterranean Evergreen Oak Forests. Springer, Berlin, pp. 121–133. Schlesinger, H.W., Gray, J.T., Gilliam, F.S., 1982. Atmospheric deposition processes and their importance as sources of nutrients in chaparral ecosystem of southern California. Water Resources Research 18, 623–629. Scudlark, J.R., Russell, K.M., Galloway, J.N., Church, T.M., Keene, W.C., 1998. Organic nitrogen in precipitation at the mid-Atlantic US coast—methods evaluation and preliminary measurements. Atmospheric Environment 32, 1719–1728. Serrasolses, I., Diego, V., Bonilla, D., 1999. Soil nitrogen dynamics. In: Roda´, F., Retana, J., Gracia, C., Bellot, J. (Eds.), Ecology of Mediterranean Evergreen Oak Forests. Springer, Berlin, pp. 223– 235. Sutton, M.A., Pitcairn, C.E.R., Fowler, D., 1993. The exchange of ammonia between the atmosphere and plant communities. Advances in Ecological Research 24, 301–393. Van Breemen, N., Mulder, J., Van Grinsven, J.J.M., 1987. Impacts of acid atmospheric deposition on woodland soils in the Netherlands. II. Nitrogen transformations. Soil Science of America Journal 51, 1634–1640. Vitousek, P.M., Aber, J.D., Howarth, R.W., Likens, G.E., Matson, P.A., Schindler, D.W., Schlesinger, W.H., Tilman, D.G., 1997. Human alteration of the global nitrogen cycle: sources and consequences. Ecological Applications 7, 737–750. Wyears, G.P., Vermuelen, A.T., Slanina, J., 1992. Measurement of dry deposition of ammonia on a forest. Environmental Pollution 75, 25–28.