ECOLOGICAL EXPOSURE AND EFFECTS OF 

AIRBORNE TOXIC CHEMICALS: AN OVERVIEW 

edited by 

1

Thomas J. Moser, 1Jerry R. Barker, and 2 David T. Tingey 

1 ManTech Environmental Technology, Inc. 2

US Environmental Protection Agency 

Environmental Research Laboratory 

200 S.W. 35th Street 

Corvallis, Oregon 97333 

NOTICE 

PB91-148460 · 

EPA/600/3-91/001 

January 1991 

The information in this document has been funded wholly (or In part) by the U.S. Environmental 

Protection Agency. It has been subjected to the Agency's peer and administrative review, and It has 

been approved for publicat:on as an EPA document Mention cl trade names or commercial products 

does not constitute endorsement of recommendation for use. 

REPRODUCED BY 

U.S. DEPARTMENT OF COMMERCE 

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SPRINGFIELD, VA 22161 
TECHNICAL REPORT ~ATA (l'laM rnd llllll'ktriotu on the rnme {on eomple 

1. fUI'OfiiT NO. 12. j:a. PB91-1.ii8460 

EPA/600/3-91/001 

•. TITLE AND IUITITLE [I. ,_1.\P'O"T DATE 

January 1991 

Ecological Exposure and Effects of jLPE"FORMINO O"OANIZATION CODE 

Airborne Toxic Chemicals: An Overview 

7. AUTMOfUSI II· ••"'o"MINO o"oANIZATION "E"O"'T No. 

Thomas J. Moser, Jerry R. Barker, and 

David T. Tingey 

ji. P'UIFO"MING O"GANIZATION NAME ANO ADOflUS 10. I'"OOflAIIot ELEMENT NO. 

1st two, NSI, ERL-Corvallis; Tingey, 

us EPA, ERL-Corvallis r1 1. eoN ""'"',, .. " .. NT No. 

12. I"ONSOfiiiNG AGENCY NAME ANO AOOfliSS 1a. TYPE OF REP'OflT AND P'EfliOO COVE fliED 

US Environmental Protection Agency Pnhl;-::ho:>n 'D 

Environmental Research Laboratory 1 •• PONSO"ING AGENCY CODE 

200 SW 35th Street EPA/600/2 

Corvallis, OR 97333 

11. I~''"LEMENTAfiiY NOTES 

1990. u.s. Environmental Protection Agency, Environmental Research 

Laboratory, Corvallis, OR. 16. Aa5Tfii-'CT 

Since the release of the Environmental Protection Agency's (EPA's) toxic release inventory (TRI) estimates for 1987, there has been a 

heightened concern over the nation's air quality. Primarily, this 

concern has been directed at human health effects in industrial-urban 

areas. The fact that many airborne chemicals pose hazards to human 

health is only one aspect of the problem. The continued deposition of 

airborne toxic chemicals pose threats to both terrestrial and aquatic 

ecosystems, this discussion is limited to terrestrial vegetation. < -

~7. ICEY WOflDS AND DOCUMENT ANALYSIS 

Ia. DlSCRIPTOflS b. IDENTIFIERS/OPEN INDIO TERMS c. COIATI field/Grou~ 

11. O>STI'IIIUTION STATEMENT t1tfrl~fiiiTY. f .. :J ('17til.eport} 21. NO. OF PAGES 

nc ass1 1e 170 

Release to Public 20. SECUfliTY CLASS (fJtU ~~ 22 ..... 1CE 

Unclassified 

&PA ,.., ... auo-t , .. 7Jl 
TABLE OF CONTENTS 

Table of Contents 

Preface 

Atmospheric Transport of Toxic Chemicals and Their Potential Impacts on Terrestrial 

Vegetation: An Overview (Thomas J. Moser, Jerry R. Barker, and David T. Tingey) 

Organochlorines and Polycyclic Aromatic Compounds In the High Arctic 

(Grag W. Patton and Terry F. Bidleman) 

Air-Sea Exchange of Hexachlorocyclohexane in tr.e Bering and Chukchi Seas 

(Daniel A. Hinckley, Terry F. Bidleman, and Clifford P. Rice) 

Trace Metal Deposition and Cycling in Forested Ecosystems (Andrew J. Friedland) 

Atmospheric Deposition of Trace Metals to Lichens and Effects of Trace Metals 

on Uchens (Thomas H. Nash Ill) 

Fate and Effects of PAHs in the Terrestrial Environment: An Overview 

(Nelson T. Edwards) 

Effects of Organic Chemicals in the Atmosphere on Terrestrial Plants 

(Jeffrey R. Foster) 

Effects of Atmospheric Pollutants on Peatlands (Noel R. Urban) 

Use of the PHYTOTOX Database to Estimate the Influence of Herbicide Drift 

on Natural Habitats in Agroecosystems (James E. Nellessen and John S. Fletcher) 

Biological Markers In Environmental Sentinels to Establish Exposure to, and 

Effects of, Atmospheric Toxicants (John F. McCarthy and Timothy J. Tschaplinski) 

Air Toxics Ecological Effects Research Recommendations (Noel R. Urban and 

Jerry R. Barker) 

References 

ii 

ii 

iii 

8 

20 

32 

40 

48 

60 

90 

102 

107 

128 

130 
EFFECTS OF ORGANIC CHEMICALS IN THE ATMOSPHERE ON TERRESTRIAL PLANTS 

INTRODUCTION 

Jeffrey R. Foster 

H~~bR~rehlnR~~e 

Butler University 

Indianapolis, IN 

Thousands of organic compounds are released into the atmosphere each year by human 

activity. Of these, the U.S. Environmental Protection Agency (EPA) has classified 328 as Priority 

P~lutants, based on scientific evidence of toxicity In humans and other animals. However, only two, 

benzene and vinyl chloride, have National Ambient Air Quality Standards. UntU recently, our 

knowledge of environmental releases of these substances was sketchy. In 1986, Congress passed the 

Emergency Planning and Community Right-t~Know Act, requiring manufacturers of Priority P~l~nts 

to report the amounts of each that they released, deliberately or accidentally, into air, water, and soil. 

The results of the first round of the Toxic Release Inventory (TRI) revealed that 865,000 metric tons of 

toxic organic compounds were released to the atmosphere in 1987 (US EPA. 1989). Small 

manufacturers ( <3,400 kgjyear) and end users (e.g., farmers applying pesticides) were not required 

to report, and no attempt was made to incorporate emissions from area sources (dry cleaners, motor 

vehicle exhausts, etc.). Thus, total emissions were probably 88Y8I8I times larger than the TRI data 

Indicate. 

Following emission, organics are dispersed widely In the environment long-distance 

transport occurs primarly In runoff, groundwater, and the atmosphere (Cohen, 1986). The latter 

pathway has spread organics to the most remote regions of the globe (Schroeder and Lane, 1988). 

Chlorinated pesticides such as DDT, lindane, dieldrin, and hexachlorobenzene (HCB) are the bestknown examples. having been detected in inorganic media and organisms as distant from emissions 

60 
as the Arctic and Antarctic (Atlas and Giam, 1981; Tanabe at al .• 1983; Gregor and Gummer, 1989). 

Long-<jistance dispersal d nonpesticide organics, Including polycyclic aromatic hydrocarbons (PAHs). 

polychlorinated biphenyls (PCBs), chlorophenols, and alkanes, has also been demonstrated (Lunde 

and Bjorseth, 19n; Heit et al., 1984; PaasMrta et al., 1985; Wickstrom and Tolonen, 1987; Gregor 

and Gummer, 1989). 

What are the consequences d this global organic contaminant load on the structure and 

function d organisms and natural ecosystems? 

PROBLEM STATEMENT 

A vast amount of literature has been accumulated on the ecotoxlcology d organics In natural 

environments. However, it is quite clear that the driving force for research has been the concern over 

their accumulation In crops for human or livestock consumption, primarily as residues following 

pesticide applications, and in accumulation via biomagnlficatlon up aquatic and terrestrial food chains 

to organisms that are consumed by humans (e.g., fish), or organisms that bring us aesthetic pleasure 

(e.g., birds). There is, comparatively, a severe dearth d information concerning the fate and effects 

d toxic organics in noncrop, terrestrial, vascular plants. 

The only comprehensive assessment d the literature on organic chemicals and terrestrial 

plants is the EPA-funded PHYTOTOX computer database (Fletcher et al., 1988). This database 

contains n,825 dose-response (effects) records gleaned from 9,700 bibliographic references published 

between 1926 and 1984. Of the 1,569 species represented In PHYTOTOX. only 417 are native 

species growing in natural habitats. The latter value Increases to 557 If old-field-succession species 

are included. Together, wlld-grOYm and old-field species constitute 4,524 records, or 5.8% d the total 

database. Thus, the existing literature is heavly oriented toward agronomic species. Pesticides 

constitute 21% d all PHYTOTOX effects records. In fact. ci the 20 most often cited chemicals, only 

two (both hormones) are not pesticides. And, cny 4.71E. d records are for native plants growing in 

natural environments. 

From the toxicological standpoint alone, I Is evident that lnsufficiet"lt data exist for EPA to 

undertake risk analyses for the effects d organics on natural terrestrial vegetation. In a:klition, litUe is 

known concerning deposition rates d organics to terrestrial ecosystems, rates d plant uptake, and 

subsequent translocation and metabolism. 

61 
The objectives d this report are to: (1) discuss mechanisms d atmospheric deposition of 

organic substances to t•arrestrial vegetation; (2) describe probable pathways d terrestrial vascular 

plant uptake d organics; (3) summarize existing literature on the effects d organics on terrestrial 

vascular plant growth and physiological processes; (4) specuate on potential, as yet unstudied, 

effects; and (5) recommend areas for future research. 

I rely heavly on the herbicide literature, which greatly exceeds that for all other organics 

combined, but make no claim to provide any but the most cursory overview. i wll consider all kinds 

d airborne organics, with the exceptions of peroxyacetyl nitrate (PAN), a lcnaNn gaseous oxidant and 

phytotoxin that has received much attention during the last 20 years (see Mudd, 1975, for a review), 

and the chlorofluorocarbons. Also, I Y'tll not review the effects d water-borne organics on aquatic 

macrophytes. 

CONCEPTUAL APPROACH 

Figure 1 shows a model of the various fluxes of airborne organic substances onto, Into, and 

through terrestrial vascular plants. Fluxes specific to the canopy are shown In detal In Figure 2. In 

the discussion that follows, I will deaf with most d these fluxes, with the exception d within-soil 

fluxes, which are beyond tt-.e scope c:l this paper. Following that, I wll describe known effects of 

organics on terrestrial plants, and finish with recommendations for future research. 

ATMOSPHERIC DEPOSITION 

Atmospheric Occurrence 

Organics enter the atmosphere primarily In gaseous or particulate form. Initial mixing In the 

air leads to partitioning between the vapor, solid, and liquid phases (Schroeder and Lane, 1988). 

Phase distributions are determined by inherent chemical properties, Including vapor pressures, Henry's 

Law constants (H. equlibrium ratio d the concentration In air to concentration In water), diffusion 

coefficients in air and water, and equilibrium constants (pK.; for weak acids) (Tucker and Preston, 

1984; Mackay et al., 1986; Bidleman, 1988; Ugocld and Pankow, 1989). Organics enter precipitation 

by dissolving directly from the vapor phase Into water droplets and by being physically scavenged as 

aerosols by falling droplets. The -washout ratio• fN; mass d organics per unit volumed rain divided 

by mass d organics per unit volume d air) describes the net effect d these processes (Bidleman, 

1988). 

62 
wet& dry deposition 

•elatiiiUIIOII 

adsorption 

wet&dry 

deposition 

translocation 

& 

detoxification 

d~~~~iion <::= ~ ---> un'fl~~ated 

degradation 

biological r:/' {!, II root uptake ~ "-.... root 

exudation 

leaching 

------ water table ---

FJGURE 1. Conceptual model d fluxes d airborne organics to and within terrestrial ecosystems. 

63 
wet 

deposition 

wastioff 

stomatal 

absorption 

cuticle 

absorption/ 

IMfsorption 

surface 

moisture 

absorption 

I resuspension 

volatilization photodecomposition 

FIGURE 2. Same as Fig. 1, but enlarged to show fluxes associated with leaves. 

64 
Compared to the ma!or air pollutants (ozone, sulfur dioxide, nitrogen dioxide, carbon 

monoxide), air concentrations of organics are not well-characterized. A great variety have been 

detected whenever and wherever they have been looked for. Urban air contains hundreds of volatile 

organic compounds (VOCs) from industrial emissions, motor vehicle exhausts, and other 

anthropogenic sources {Atlas and Giam, 1988; Bruckmann et al., 1988; Shah and Singh, 1988; 

Edgerton et al., 1989). Some VOCs are secondary products of the photochemical reactions that 

produce smog (e.g., nltrophenols [Nojima and Kanno, 1977)). These VOCs are transported downwind 

to rural regions (Rice et al., 1986; Rippen et al., 1987; Shah and Singh, 1988; Schroeder and Lane, 

1988). Reverse transport also occurs; although having highest concentrations in agricultural areas, 

chlorinated pesticides are regularly found in urban air (Abbott et al., 1966; Stanley et al., 1971; 

Kaushik et al., 1987; Bruckmann et al., 1988). Natural vegetation, especially trees. puts out substantial 

quantities of VOCs. primarily monoterpenes, which mix with anthropogenic VOCs (Roberts et al., 1985; 

Trainer et al., 1987; Kreuzig et al., 1988; Petersson, 1988}. Table 1 summarizes the range of 

concentrations of selected organics measured in air and rain. 

TABLE 1. CONCENTRATIONS OF SOME PESTICIDES, PCBS, PAHS AND OTHER VOLATILE 

ORGANIC COMPOUNDS IN AIR AND WATER, UNITED STATES• 

Compound Air (ngjm~ Rain {ngjL) 

Hexachlorobenzene 0.13~.29 0.40-1.0 

a.-j..,...HCH 0.30-5.4 6.4-37 

DDT/DOE 0.02~.33 0.50-4.3 

Chlordane 0.04-1.3 trace-2.3 

Dieldrin 0.02~.08 0.8-2.0 

Toxaphene 0.6-13.1 7.31-59 

PCBs 0.06-9.3 0.10-200 

Pthalate esters 1.0-290 >200 

Benzo(a) pyrene 1.3-500 2.2-7.3 

Formaldehyde 2375-12250 b <0.01-1.1 

Acetaldehyde o-9900b <0.04-1.1 

Benzene 1950-10700 b e 

Phenol 39-430b <1G-14000 

Benzo(a) pyrene 0.01~1 

• Sources: Pearson (1982), Nriagu and Simmons (1984), Giam et al. {1984), Chapman et al. (1986), 

Atlas and Giam (1988), Shah and Singh {1988), Leuenberger et al. (1988), Dulnker and Bouchertall 

(1989), Sawicki et al. (1960) 

o lower jupper quartiles 

e no data availa~e 

65 
Wet Deoositlon 

The flux cl organics to terrestrial ecosystems in precipitation is simply the volume-weighted 

product cl the concentration In solution and precipitation volume, summed over the tirne period of 

interest. A. growing literature reports concentrations d organic compounds In rain. Most regularly 

and widely detected are organic acids, especiaJiy aldehydes such as formate and acetate (Chapman 

et at., 1986 and references therein; Gaffney et al., 1987). Chlorinated pesticides, PCBs, phenols, and 

n-alkanes have all been reported In rain (Eisenreich et al., 1981; Farmer and Wade, 1986; Agarwal et 

al., 1987; Gaffney et al., 1987; Jones et al., 1989; Leuenberger at al., 1988; Atlas and Giam, 1988). 

FfNI estimates of wet deposition flux for organics exist. Farmer and Wade (1986) observed 

22 to 670 JJQ/rn2/day for C,2 - C32 hydrocarbon fluxes at Norfolk, VA. An estimate for bulk 

precipitation, which incorporates some aerosols, was 40 and 20 pgfm2

/y for PCBs and lindane, 

respectively, at Paris, France (Chevreuil et al., 1989). Precipitation falling onto the Great Lakes has 

been analyzed for PCBs and pesticides for some years. Deposition rates are 0.6 to 9.6 pgfm2

/y for 

chlorinated hydrocarbons, 424 pgfm2

/Y for PAHs, and 120 pgjm2

jy for PCBs (Nriagu and Simmons, 

1984). 

Drv Deoosition 

Dry deposition of atmospheric substances onto natural vegetation is described by the 

equation 

where F = flux density, C. = atmospheric concentration at a reference height above the vegetation, 

and V d = deposition velocity. Deposition velocity has four components: 

where r. = bulk aerodynamic resistance, r b = leaf or sol boundary layer resistance, r. = stomatal 

resistance, and r c = chemical resistance. 

Aerodynamic resistance represents the turbulent downward movement of parcels c1 air, 

carrying entrained gases and particles. It is affected by wind speed, the surface area, and vertical 

distribution of canopy components (foliage, branches, etc.). Boundary layer resistances are related to 

wind speed and wind direction, and to collecting surface size and geometry. Stomatal resistance is 

66 
directly proportional to stomatal aperture, which is under physiological control by plants. Chemical 

resistance comes Into play when the concentration d the compound at the absorbing surface is not 

zero (i.e., It is not quantitatively destroyed or sequestered upon deposition). In this case, the 

concentration gradient driving flux wll be less than C •. 

Deposition d gaseous organics occurs primarly to dry sol and vegetation surfaces and Is 

controlled largely by salubRity in the waxy cuticles d plant surfaces and vapor-sol particle partitioning. 

Absorption d gaseous organics Into water is Important to wet sols, on plant surfaces following rain or 

during fog and dew events, In permanent wetlands, and In riparian areas during flooding. Deposition 

of organics In aerosol form Is achieved by the same mechanisms as for aerosols In general: 

sedimentation, diffusion, and impaction. Sedimentation is most Important for large aerosols, diffusion 

for small aerosols, and Impaction for Intermediate-sized aerosols (Garland et al., 1988). 

Dry deposition of organics has been measured directly by accumulation on surrogate 

surfaces, Including silicon oil-coated nylon screens (Sodergren, 1972); glass plates sprayed with 

mineral oil or water (McClure and Lagrange, 1977; Murphy, 1984); stainless steel funnels and sinks 

(Farmer and Wade, 1986; Chevreuil et al., 1989); and pans filled with water, ethylene glycol-water, or 

glycerin-water (Bidleman and Christensen, 1979; Christensen et al., 1979). Observed fluxes have been 

In the range 10-3 to 102 J.Ig/m2

jday for PCBs and various pesticides, yielding V d values of less than 

0.01 to more than 5 cmjs. Confidence in these fluxes is not high because the efficiency collection of 

surrogate surfaces may vary substantially from that of natural surfaces. Bidlernan and Christensen 

(1979) found that V d for PCBs, chlordane, and DDT varied geographically and by one to two orders of 

magnitude from day-to-day. In addition, aerosol and gaseous deposition cannot be distinguished, and 

volatilization of adsorbed gases Is not accounted for. Deposition velocities measured by accumulation 

have been used to infer dry deposition to the Great Lakes from known concentrations of organics in 

air (Eisenreich et al., 1981). 

Fogs. Clouds. and Dew 

Cioud water and fog water contain higher concentrations d organic acids and much higher 

concentrations d pesticides than does rain (Glotfelty et al., 1987; lgawa et al., 1989; Munger et al., 

1989; Schaefer et al., 1989). Organic acids have also been reported from dew (Mulawa et al., 1986). 

Interception of cloud water and d advective (seacoast) fog water Is modeled as a dry 

deposition process, with cloud and fog droplets acting like large aerosols (Lovett, 1984). Deposition 

of radiation (valley) fog water Is simDar to precipitation O.e., gravitational settling d water droplets). 

Dews form by condensation of water vapor onto surfaces whose temperature drops to the dewpoint; 

67 
deposition rates can be predicted using the Penman-Monteith equation. Organics enter dews by 

gaseous absorption and aerosol sedimentation. There appear to be no deposition estimates for 

organics in fogs, clouds, and dews. 

Total Atmospheric Deoosition 

Accumulation rates of persistent (slow to degrade) organics in sediments yield an estimate of 

net atmospheric deposition (i.e., not accounting for volatDization). For example, Heit et al. (1981, 

1988) and Gschwend and Hites (1981) measured PAHs In lacustrine and marine sediment cores in the 

northeastern United States. Ruxes in recent decades ranged from 36 to 4,870 1Jg/m2

fy, with greater 

fluxes near urban areas. Jones et al., (1989) measured PAHs in archrved, dry soil samples collected 

since 1843 in rural England. Mean deposition rates (1880-1980) were 4,560 1Jg/m2

jy. Wickstrom and 

Tolonen (1987) found that PAH fluxes into sediments of small Finnish lakes had increased 

substantially since the earty nineteenth century. 

Swackhamer and Armstrong (1986) estimated net atmospheric deposition of 21Jg/m2

/y of 

PCBs in recent decades to several small seepage lakes in Wisconsin. Rapaport and Eisenreich (1988) 

measured PCBs, HCB, DDT, and toxaphene accumulation in peat bog sediments. Maximum 

deposition rates in the eastern United States occurred between 1960 and 1978, with reduced rates 

since 1980. 

The only estimate that could be found for recent atmospheric deposition of an organic to a 

terrestrial ecosystem was that of Matzner (1984) for German beech and spruce forests. Assuming no 

root uptake or exchange between surface deposits and canopy tissues, he estimated total PAH fluxes 

of 0.25 to 0.69 Jlgfm2

fy. 

A regional application of mass balance was applied to the Great lakes by the International 

Joint Commission (summarized in Elder et al., 1988). Measurements of PCBs and pesticide 

concentrations in dated sediments yielded values for long-term inputs. When known contributions of 

tributaries, runoff, and point sources were subtracted from these values, and outflows added, the 

difference was ascribable to atmospheric deposition. The atmosphere accounts for 60% of PCB 

inputs and 78% of benzo[a]pyrene inputs to the Great Lakes. 

68 
FLUXES OUTSIDE OF PLANTS 

Fates In Soil 

Once deposited, the disappearance a organics in sols usually follows first-order kinetics, 

provided no additional input occurs, and plant uptake Is either constant or absent (Ryan et al., 1988). 

Factors contributing to the disappearance a organics Include volatlization, photodecomposttion, 

adsorption to sol particles, chemical and biological (microbial) degradation, unsaturated and saturated 

flow, and leaching to deeper sol layers (Figure 1 ). Consideration a these processes (except 

volatilization) Is beyond the scope a this paper. However, modeling a disappearance rates Is useful 

In determining the likelihood a organics to accum&.Nte to concentrations at which plant root uptake 

will become significant (Ryan et al., 1988). 

Volatilization 

In general, volatilization rates of organics from plant canopies Increase with increasing vapor 

pressures and decreasing Henry's law constants (H). Foliar volatUization rates do not fit simple firstorder kinetics because of diurnal variability in environmental conditions (solar radiation, wind speed, 

etc.), and plant uptake (Willis et al., 1986). There is no simple relationship between volatilization rates 

of organics from soils and either vapor pressure or H values; such factors as adsorption coefficients 

to soil particles and soil moisture content are also Important (Kizer et al., 1979; Mackay et al., 1986; 

Spencer et al., 1988). Measured volatilization losses of herbicides In agricultural fields show great 

variability; 4.5% per year for dieldrin, but less than 1% per year for carbofuran (Caro et al., 1976); 

80% per year for PCBs (Moza et al., 1979); 8, 24, and 34% over a six-(iay period for toxaphene, 

parathion, and fenvalerate, respectively (W~Iis et al., 1986): less than 1% over seven days for diclofop 

(Smith et al., 1986); and 21% over five clays for 2,4-0 (Grover et al., 1985). 

A substantial proportion of herbicides volatDized from soils can be redeposited on crop 

canopies, especially DDT, PCBs, and the nltrophenol dlnoseb (Parker, 1966; Prendeville, 1968; Beall 

and Nash, 1971; Kaufman, 1976; Moza et al., 1979; Fries and Marrow, 1981; Bacci and Gaggl, 1985). 

Topp et al. (1986) found that the proportion of sol-applied organic chemical uptake by barley foliage 

increased as the volatilized fraction of applied chemical increased. 

Wash-Off and Resuspension 

Wash-off occurs when rain physically dislodges particles or carries away dissolved gases and 

particulates from plant surfaces; these are then deposited to the sol in throughfall (Figures 1 and 2). 

The highest proportion of wash-off to amount applied is for pesticides with high solubDity in water, for 

69 
those applied in particulate form, and when heavy or repeated rains occur shortly after application. 

Time-course studies d individual rain events reveal that wastHlff Is rapid at first, then diminishes 

asymptotically toward zero. Wash-off increases with Increasing rain amount, up to some saturating 

intensity Qsensea and Jones, 1971; Steffens and Weineke, 1975; Cohen and Steinmetz, 1986; Willis et 

al., 1986). 

Resuspenslon Is the raentrainment r:i particles deposited on plant surfaces in moving air 

(Figure 2}. Particle adhesion and windspeed are the primary factors lnlluenclng this process 

(Nicholson, 1988). Soil partides may be entrained In the atmosphere (wind erosion) or ejected into 

the air by raindrop splash (Chamber1aln, 1975), from whence they may be redeposited on plant 

surfaces (Pinder and McLeod, 1988). 

FlUXES INVOLVING PLANTS 

Organics must penetrate plant tissues before they can exert physiological effects. Thus, 

aerosol deposition and nonstomatal deposition d gases are appropriately considered as separate 

processes from plant uptake. In the case r:i stomatal absorption, deposition and uptake are 

equivalent because the sink Is the mesophyll. 

Uptake rates are determined by external concentration, concentration in plant tissues. and 

resistances to uptake. For shoots, the most Important physical barrier Is the cuticle. For roots, the 

cell membranes of the endodermis are a significant barrier. Following uptake, if the substance Is not 

translocated or metabolized to other forms. It may accumulate in tissues, reducing the concentration 

gradient across the plant surface, and thc;s reducing flux. Translocation to other parts of the plant, or 

metabolic conversion to other compounds, will lower Internal concentratiori near the sites r:i uptake, 

maintaining the concentration gradient. 

Foliar Uotake of Vaoor-Phase Organics 

Radioisotope studies with carbon-14 (

14C) labeled organics have demonstrated that In most 

plant species. uptake r:i PCBs Is primarily by leaf vapor absorption, even when the PCBs are applied 

to the sol (Nash and Beall, 1970; Iwata and Gunther, 1976; Weber and Mrozek, 1979; Bacci and 

Gaggi, 1985). Nash and Beall (1970) and Beall and Nash (1971) found the same to be true of 

soybean plants exposed to sol-applied DDT, but sol volatKization was less Important for dieldrin, 

aldrin. and heptachlor. PCBs accumulate in both herbaceous and woody species in proportion to 

PCB concentration in the surrounding atmosphere (Buckley, 1982). Soybeans absorbed 1

"C70 
anthracene vapors from air (Edwards et al., 1982). AccumUation c:i PCSs, DOT, penta-chlorophenol 

(PCP), nitrophenols, dioxins, furans, and DGohexachlorocydohexane (CG-HCH) also occurs In conifer 

foliage (Eriksson et al., 1989; Hinkel et al., 1989, Reischl at al., 1989a). 

Gaseous hydrocarbons are raadYy absorbed by plant foliage. When tomato. barley. and 

carrot plants were exposed to hydrocarbon vapors (2 x 10-4M). foliar damage symptoms appeared 

rapidly (Currier, 1951; Currier and Peoples. 1954). Frank and Frank (1989} exposed Norway spruce 

saplings to tetrachloroethane in glass chambers. The concentration c:i tetrachloroethane In the 

needles (Cn) Increased linearty with Increasing concentration In the chamber air. At Cn < 5 pgfcm~ 

the partitioning between air and needles was greater than when en> 10 pgfcm~ suggesting surface 

adsorption, folltwied by cuticular absorption when adsorption capacity was saturated. 

Wolverton et al. (1984) reported that house plants absorbed formaldehyde from the air. 

However. additional experiments by Godish and Guindon (1989) established that the absorption was 

by potting soil rather than the plants. Bioconcentration of vapor -phase organics in leaves Is closely 

correlated with their octanol-water partition coeffteients [(K_ (Figure 3); Travis and Hattemer-Frey, 

1988; Reisch! et al., 1989b]. 

Foliar Uptake of Uauid-Pbase Organics 

Penetration of the cuticle by herbicides applied in solutions is positively related to their K_ 

values (Shafer and Schonherr, 1985) and Is usually linearty related to herbicide concentration 

(Bukovac, 1976). Uptake is generally greater to adaxial than to abaxial leaf surfaces, apparently 

because of the higher density of stomata and greater pubescence that occur on adaxial surfaces 

(Sargent and Blackman, 1962; Bukovac, 1976). Trichome bases and guard cells are preferred 

locations of uptake. Mass movement through stomata occurs If the organic, or its solvent, are 

lipophilic and have low surface tension (Baker, 1970; Bukovac, 1976). Most studies of foliar liquid 

uptake by herbicides, using 1

"C labeling, have found substantial uptake (Bukovac, 1976; Steffens and 

Wieneke, 1975; van Auken and Hulse, 1979), but 2.~ichlorophenol and dioxin do not appear to 

penetrate the cuticle Osensee and Jones, 1971). 

When Currier (1951) and Currier and Peoples (1954) exposed plants to pure hydrocarbon 

liquids, development c:i foliar damage was rapid, Implying fast uptake across the cuticle. In a 

literature revie'N on hydrocarbon ol elfects on plants. Baker (1970) concluded that those ols 

containing mostly small-molerue. IOIN-volatllty hydrocarbons with low viscosity are most phytotoxic; 

the hydrocarbon constituents readly diffuse across cuticles and cell membranes, evaporative losses 

are low when applied to sols or foliage, and penetration through stomata occurs. 

71 
6~----------------------------------------------~ 

5 

4 

, , , , 

, 

a-HCH .'J" 

,4 , " lindane 

4 

, , 

, , , , , 

, , , , , 

5 

LOG Kow 

,' , , , , , p,p'-DDT , , ' 

. ,' 

p,p'-DD~ - , ' , , ePCBs , , , , , 

6 7 

FIGURE 3. The relationship between bloconcentration of organics in plant tissues (B~; ratio of 

concentration In plant to concentration In air) and the octanol-water partition coefficient (K_) (after 

Travis and Hattemer-Frey, 1988). 

72 
Foliar Uotake of Solid-Phase Organics 

Organics are frequently adsorbed to the surfaces d atmospheric aerosols, but no studies of 

particle-associated organic uptake by plants have been conducted. However, cuticular absorption of 

lipophilic organics Is probably much faster from the vapor phase than from the solid phase (Reisch! et 

al., 1989b}. 

Root Uptake 

Manr different herbicides are absorbed by the roots d crop plants (Uchtenstein and Schultz. 

1965; Lichtenstein et al., 1967; Shone and Wood 1974; Bukovac 1976; Hlton 3l al .• 1976; Briggs et 

al., 1982; WICkliff et al., 1984; McFarlane and Wickliff, 1985; Scheunert at al., 1986; Topp et al., 1986; 

McFarlane et al. 1987a,b). A linear relationship Is often observed between soil or growth solution 

concentrations and rates of uptake, at least when concentrations are low. At higher concentrations, 

uptake may be saturated or even reduced (Aetcher et al., ~990), perhaps due to direct toxicity to the 

roots. If the original herbicide Is not very soluble, root uptake rates may instead be related to 

concentrations of soluble sol metabolites (Scheunert et al., 1986). 

Root uptake has been documented for the insecticides dieldrin, endrin, aldrin, and lindane, but 

does not occur for DDT and heptachlor (Ha~ and Sans, 1967; Lichtenstein et al., 1967; Nash et al., 

1970). Root uptake has also been observed for dioxins, benzene and substituted benzenes, phenol 

and substituted phenols, ethanol, naphthol, acetate, tiaryl phosphate esters, pthalate esters, organoborates, and trinitrotoluene (Harley and Beevers, 1963; Isensee and Jones, 1971; Kaufman, 1976; 

Suzuki et al., 1977; Moza et al., 1979; Mrozek and Leidy, 1981; Shea et al., 1982; Edwards, 1983, 

1986; Casterline et al., 1985; McFarlane and Wackliff, 1985; Scheunert et ai., 1985, 1986, 1989; 

Facchetti et al., 1986; Palazzo and Leggett, 1986; Sacchi et al., 1956; Topp et al., 1986; McFarlane et 

al., 1987a,b; Adriano et al., 1988; Aranda et al., 1989; Krstich and Schwarz, 1989; Retcher et al., 1990; 

McFarlane et al., 1990; O'Connor et al., 1990). Depending on the species, plant roots may or may 

not absorb PCBs and PAHs {Edwards 1983, 1986; O'Connor et al., 1990). 

During root uptake, organics move by diffusion In the apoplastic water d the cortex; then, at 

the Casparian strip, cross the cell membranes c:A the endodennls to reach the xylem. Shone and 

Wood (1974), Briggs et al. {1982, 1983), Topp et al. (1986), and McCrady et al. {1987) found a 

positive, l~lnear relationship between the root concentration factor (RCF; ratio of root tissue 

concentration to external solution concentration) and K_ for various organics. HO'Never, In a 

comparison of nitrobenzene absorption by eight plant species, RCF varied considerably around the 

73 
predicted value from Briggs et al. (1962), emphasizing the importance d species-specific 

characteristics (McFar1ane et al., 1990). 

Bark Uptake 

Penetration d the bark d woody plants by herbicides has been reported (Bukavac, 1976). 

Meredith and Hites (1987) found a variety d PCB congeners lri the bark of black walnut and tUiiJr 

poplar trees growing near a PCB-contaminated landfill. Very kM' PCB concentrations In the wood, 

much higher PCB concentrations in inner than in outer bark. and a positive correlation between 

congener K_ and accumt.Ation in outer bark suggested that uptake occurred from the atmosphere 

into the suberin d cork cells. 

Translocation 

When 14 C-labeled herbicides are taken up by roots, most move upward in the transpiration 

stream, but some remain in the cortex, apparently unable to cross the endodermis (Hay, 1976; Hilton 

et al., 1976). The proportion translocated varies according to length cl exposure, plant species, and 

type of herbicide (McFartane et al., 1987b). What is translocated may be found fairty uniformly in all 

the leaves or may be concentrated in young leaves and growing tips (Hay, 1976). Dry soils reduce 

rates of translocation, presumably due to lesser transpiration and a lower carbohydrate supply for 

phloem transport. 

Translocation of root-absorbed, non-herbicide organics has been demonstrated for aldrin, 

lindane, bromacn, triaryl phosphate esters, phenol and chlorophenols, dioxins, pthalate esters, PCBs, 

trinitiotoluene, and nitrobenzenes (Uchtenstein et al., 1967; Isensee and Jones, 1971; Mrozek and 

Leigy, 1981; Shea et al., 1982; Castertine et al., 1985; Edwards, 1986; Palazzo and leggett, 1986; 

Sacchi et al., 1986; McFartane et al., 1987a,b, 1990; Aranda et al., 1989; Aetcher et al., 1990). Rootand leaf-absorbed hydrocarbon oHs are transported acropetally and baslpetally, respectively (Buckley, 

1982). However, there is little translocation d root-absorbed PCBs, PAHs, and substituted phenols 

(Hilton et aJ., 1976; Kaufman, 1976; DOrr, 1970; Buckley, 1962; Scheunert at aJ., 1989), nor do root- or 

leaf-absorbed dioxins and furans appear to be translocated (Reischl et al., 1989b). 

There Is considerable evidence for a positive relationship between transpiration rate and root 

uptake (see Figure 4) (Bukavac, 1976; van Oorschot. 1976; Mcfarlane at al., 1987b). However, 

Shone and Wood (1976) discovered that neither external concemation nor transpiration rate had 

much effect on uptake and translocation d triazine herbicides by radish. 

74 
-Cl 

:1 

-

1.6~------------------------------------------------~~ 

1.2 

0.4 

Exposure Time: 

• 3d8)'S 

0 Sdays 

A 7days 

6 10days 

atrazine in wheat 

o.ok:~--------~------------_.--------------------------~ 0.0 0.4 0.8 1.2 1.6 

Transpiration x Solution Concentration (pg) 

FIGURE 4. The relationship between total uptake of atrazine from solution by wheat and the 

expected uptake by mass flow through the xytem (after Walker, 1972). 

75 
Briggs et al. (1983) observed a positive, log-linear relationship between stem concentration 

factor (SCF; calculated In same way as RCF) and herbicide K_. It required 24 to 48 h for SCF to 

stabilize, with longer equllbration times for higher K_. They postuated that stable SCF represented 

abalance between upward transport eX root-absorbed compounds and their absorption or degradation 

by stem tissues. 

The herbicide transpiration stream concentration factor (J'SCF; ratio eX cor.ce1 lbation in 

transpiration stream to concentration In external solution) In barley plants was usually less than unity, 

implying passive uptake eX most herbicides, but greater than unity for 2,4-0. Implying active uptake 

(Shone et al., 1973; Shone and Wood. 1974). The mltrobenzene TSCF was consistently less than one 

In eight plant species (McFartane et al .• 1990). The TSCF was nat related In any obvious way to 

RCF. However, Shone et al. (1974) demonstrated that the TSCF was correlated w!th the most readily 

diffusible herbicide fraction In barley roots duted with water following a period eX root uptake. This 

fraction consisted of the most lipophilic (high K_) herbicides. 

Briggs et al. (1982} observed a bell-shaped relationship between TSCF and K_ln barley 

(Figure 5). They hypothesized that for K_ < 1.8, transport is limited by diffusion across the 

endodermal membrane; whereas, for K_ > 1.8, transport is limited by the rate eX transport from the 

roots to the tops. McCrady et al. (1987) demonstrated that sorption to the xylem tissue was the main 

limitation on transport rates of organics through excised soybean stems. McFartane et al. (1987b) 

found differential root uptake and translocation to the shoot eX soybean by bromacl, nitrobenzene, 

and phenol, although the three chemicals have simUar K_ values. Given the scatter in Briggs et al. 

{1982) relationship, this result is not surprising. Boersma et al. (1968) have developed a model of 

organic compound transport In the xylem and phloem. This model is most sensitive to K_. rates of 

detoxification, membrane permeabUity, and the volume of xylem and phloem in various organs. 

Generally, the bulk of foliar-applied organics Is not translocated; that portion that is partitioned 

to the stem, little goes to the roots (Hay, 1976; Shone and Wood, 1976; Weber and Mrozek, 1979). 

Foliar-applied ureas, triazlnes, DNOC, PCP, and hexachlorophene are not translocated (Fogg, 1948; 

Hay. 1976; Kaufman, 1976; van Auken and Huse. 1979). 

Detoxification 

Plants detoxify herbicides prlmarly by metabolic degradation and to a lesser extent by 

adsorption to biomoi8CUes. Adsorption to lecithin and cell membranes has been demonstrated for 

2,4-0 and 2,4,5-T (Wain and Smith, 1976). The metabolic fates c:l many different herbicides have 

been aucidated by 14C labeling studies (Naylor. 1976). In most cases, metabolism converts the 

76 
1.0 

• barley 

0.8 -

• 

0.6 ~ 

~ • _ ... -.--- ' .... • 

u 

\1'1 ,._, .. • ... ~ • ... .... 

0.4 , , , ' ' , • ' , 

• , , ' , ' ' 

0.2 

. , • , \ 

e,' \ , , • \ 

\ 

• , 

\ 

. 

0.0 _, 0 1 2 3 4 5 

LOG Kow 

FIGURE 5. The relationship between transpiration stream concentration factor (TSCF; ratio of 

concentration in transpiration stream:concentration In nutrient solution) c:l various organics and 

octanol-water partition coefficients (K_) In barley (after Briggs et al., 1982). 

77 
phytotoxic herbicide to nonphytotoxic or less phytotoxic forms; however, the reverse is sometimes 

true (e.g., amitrole, bipyridyliums, 2,4-D precursors, DNOC esters) (Wain and Smith, 1976; Menn and 

Still, 19n). Generally, detoxifteation proceeds in two steps: (1) an oxidizing, reducing, or hydrolyzing 

reaction that modifies the organic toxin by adding reactive groups; and (2) a subsequent conjugation 

reaction that further reduces toxicity and yields a metabolite In a form suitable for storage or transport 

(Kaufman, 1976; Menn and Stll, 1977). 

The metabolic fates of most nonherbicide organics in plants are not well understood. 

Degradation of P AHs, pthalate esters, 4-nitrophenol dichlorobiphenyts, nitrobenzenes and 

trinitrotoluene have been reported (000. 1970; Moza et al., 1979; Edwards, 1986; Palazzo and 

Leggett. 1986; McFar1ane et al., 1987a, 1990; Aetcher et at., 1990; Preiss et al., 1989; Schmitzer et al., 

1988). but the phytotoxicity of the metabolites is unknown. 

One major mechanism of plant detoxification cl organics Is conjugation to glutathione. For 

example, com, sorghum, and sugarcane, which possess this metabolic capability, resist atrazine; 

whereas broadleaf weeds and grasses, which do not have this capability, are susceptible to atrazine. 

Schroder and Rennenberg (1989) reported glutathione conjugation of pentachloronitrobenzene (often 

sprayed on crops to protect them from pathogens). but not of atrazine, by Norway spruce. 

Plant Efflux 

Efflux of organics from plants occurs in three ways: (1) foliar efflux or •leaching: (2) passive 

loss by transpiration, and {3) root exudation (Rovira, 1969; Tukey, 1970). Radiocarbon studies have 

shown that root exudation of herbicides is an energy-requiring process, occurs in the zone of active 

root elongation, and involves readily translocatable and little-metabolized herbicides (Bukovac, 1976). 

The extent to which organics taken up by plants may be subsequently lost by foliar efflux apparently 

has not been studied. 

Siddaramappa and Watanabe (1979) exposed rice plants to 14C-carbofuran in nutrient 

solutions sealed at the root collar to prevent volatUized carbofuran from reaching the leaves. 

Transpired carbofuran was trapped in ethylene rjycol in shoot chambers. Over a 1(kiay period, 9 to 

17% of the absorbed radioactivity was lost via transpiration, apparently carbofuran was carried 

passively along in xylem water. Higher transpiration rates resulted in more carbofuran recovered from 

the air. These authors also found significant radioactivity In guttation water. McFarlane et al. (1990) 

found that 10-40% of not -absorbed ~ .. C-nitrobenzene volatilized from leaves d crops and woody 

plants. 

78 
KNOWN EFFECTS ON PLANTS 

Visual lniurv and Growth 

As with plant uptake, there is a close association between K_ values of organics and their 

inhibitory effects on plant germination and growth. A good model that fits a wide range of plant data 

is a nonlinear equation where toxicity increases linearty up to a critical valued K_. then falls off at 

higher K_ (Hansch et al., 1989). 

Herbicides 

The mechanisms by which herbicides Influence growth Include Inhibition d mitosis, inhibition 

or stimulation d tissue enlargement and elongation, and alteration d patterns d tissue differentiation 

(Cartwright, 1976). Herbicides may also alter dry mass distribution among plant organs, prevent 

poliination, and cause abscission of leaves and other plant parts (Addicott, 1976; Ratsch et al., 1986); 

and, in the case of auxin~ike herbicides such as 2,4-0, abnormal growth of various organs (van Andel 

et al., 1976). Effects may be manifested at all stages of growth from germination to reproduction. 

For example, Sund and Nomura {1963) observed that some herbicides were most effective in 

inhibiting germination, others at Inhibiting seedling growth, and still others at inhibiting growth of 

juvenile plants. Several herbicides, Including dinoseb, PCP, 2,4-0 and 2,4,5-T were inhibitory at two 

or more of these life stages. 

Substituted phenols vary widely In their effects on germination and seedling growth (Sund and 

Nomura, 1963; Amer and Ali, 1968; Shea et al., 1983). Root growth Is stimulated by 2,4-dinitrophenol 

(DNP) at low concentrations and is Inhibited at high concentrations, probably because it is a 

respiratory uncoupler (Shea et al., 1983). Chlorosis, wilting, and necrosis of peas Is caused by 

2-nitrophenol and 2,4-dichlorophenol (Amer and Ali, 1968). 

Van Haut and Prinz {1979) studied the effects of a number of organic vapors on the growth of 

several crop species. Over a wide ranged concentrations (0.1-7.0 mg m -s). relative growth 

reductions were positive log~inear for all chemicals and species studied. Growth reductions relative 

to those caused by sulfur dioxide at equivalent concentrations varied between plant species and were 

usually higher at lower concentrations. The most phytotoxic substance was ethylene (up to eight 

times the growth reduction of sulfur dioxide), followed by formaldehyde and acetic acid. 

Dimethylformamide and methanol were less phytotoxic than StJfur dioxide. Dichloromethane, toluene, 

trichloroethylene, acetone, and xylene caused little or no growth reduction at concentrations up to 

60-160 mg/rn3

• 

79 
Surfactants are often used to improve "wettability'" c:i foliar-applied herbicides. These 

substances may In themselves Increase, reduce, or have no effect on the growth of crops 

(Uchtenstein et al., 1967; Singh and Orsenlgo, 1984). 

Other Organics 

Toxicity c:i pure hydrocarbon sprays applied to plants Increases In the order: straight-chain 

paraffins < napthenes and alkenes < aromatics (Crafts and Reiber, 1948). When bartey, tomatoes, 

and carrots were exposed to hydrocarbon vapors, toxicity increased In the order: benzene < toluene 

< xylene < trimethylbenzene (Currier, 1951}. The symptoms started with increased foliage odor, 

followed by the appearance c:i dark areas on the foliage, v.-iting, and, at higher concentrations, death. 

Shoot d~.l mass gain was positively correlated with vapor concentration for sublethal doses. When 

these same compounds were sprayed on foliage as pure liquids, the same order c:i toxicity was 

observed, but the differences among hydrocarbons were greater, apparently because of differences in 

volatilization (vapor pressure decreasing in reverse order to toxicity). Currier and Peoples (1954) 

repeated these experiments, finding that toxicity Increased in the order: hexene < hexane < 

cyclohexane < cyclohexene < benzene. 

Crude oils, their refined products, and coal liquid cause visual Injury and reduce growth of a 

variety of weed and crop species (Baker, 1970; Blankenship and I.Jlrson, 1978; Wamer et al., 1984). 

Toxicity is due to the mixture of hydrocarbons In these liquids, including alkanes, alkenes, napthenes, 

naphthalenes, phenols, and aromatics. Many hydrocarbons are known to be toxic in the presence of 

ultraviolet radiation (Larson and Berenbaum, 1988). For example, the PAH fluoranthene causes foliar 

injury under UV light (Zweig and Nachtigal!, 1975). Creosote, a widely used wood preservative, 

contains a variety of PAHs and other aromatics. Concentrations of 18 to 34 mgjL reduced root 

growth of onion by 50% (Sundstrom et al., 1986). 

Aroclor 1254, a mixture of PCBs, Inhibited height growth and fresh mass Increase of both 

soybean shoots and roots; at higher concentrations, newly formed leaves showed abnormal curting 

(Weber and Mrozek, 1979). Pthalate esters caused reduced growth and chlorosis of ll8"N leaves and 

growing tips of tobacco and com (Buta, 1975; Shea et at., 1982). Tributyl phosphate reduced root 

growth c:i rice, radish, and soybean at concentrations of 10 to 100 ~Jg/g sol (Muir, 1984). 

Trinitrotoluene (TNT) lowered root and shoot growth a nutsedge (Palazzo and Leggett, 1986). 

Sodium tetraphenytborate (NaTPB) reduced growth c:i loblolly pine seedlings because of Increased 

plant boron uptake to toxic levels (Kaplan et al., 1988). NaTPB's major degradation product, 

diphenytborlc acid (DPBA), did not affect growth. However, both NaTPB and DPBA, as well as 

biphenyt, another breakdown product, reduced growth c:i sorghum, with NaTPB having the greater 

Influence (Adriano et al.. 1988). 

80 
Many substituted phenols have auxin-like activity. Harper and Wain {1969) found that the 

greatest growth promotion in standard auxin activity tests was produced by phenols with substitution 

at both ortho-positions. Uke phenols, PAHs can stimulate plant growth (Grit and Nowak, 1966). 

Several alcohols promoted growth c:X excised wheat roots In the light, whereas others had no effect 

(Gudjonsdottir and Burstrom, 1962). Hexadecanol and docosanol decreased growth c:X tobacco 

{Bourget and Parups, 1963). Organic acids and aromatic alcohols. aldehydes, and carboxylic acids 

inhibited lettuce germination (Mayer and Evenari, 1953; Reynolds, 1978). The Insecticides parathion, 

diazinon, and lindane significantly reduced root growth, but not shoot growth, d peas; whereas, 

simazine reduced both root and shoot growth (Lichtenstein et al., 1967). Vapor-phase aldehydes at 

concentrations > 0.2 ppm caused foliar Injury to petunias (Brennan et al., 1964). 

Gas Exchange and Water Relations 

Herbicides 

For most herbicides studied to date, stimulation, Inhibition, or no effect on dark respiration 

have been reported (van Oorschot, 1976). In some cases, lower concentrations stimulate; whereas, 

higher concentrations Inhibit, a typical response for oxidative uncouplers such as DNP (Shea et al., 

1983). Respiration may also be affected dtfferential1y by the same herbicide in different plant organs. 

Inhibition of net photosynthesis has been observed in a wide variety of herbicides when plants 

are exposed to herbicides In sons or nutrient solutions, or applied to foliage (van Oorschot, 1976). 

With the exception of acylamides, virtually all herbicides tested to date inhibit net photosynthesis and, 

when a range of concentrations Is studied, Inhibition Increases with increasing concentration. Unlike 

respiration, stimulation of net photosynthesis has not been observed (van Oorshot, 1976). 

All herbicides that reduce net photosynthesis reduce transpiration as welt, such as ONP 

(Barber and Koontz, 1963; Pemadesa and Korelege; 1977). This phenomenon raises the question as 

to whether or not photosynthetic inhibition is primarily due to stomatal closure, or to nonstomatal 

(biochemical) effects. Nonstomatal inhibition can be represented as a •mesophyll resistance• (r .J In 

series with r band r ,. lmbamba and Moss {1971} found that r"' for C02 in com exposed to atrazine 

increased without an accompanying change in r,. Most evidence for the relative roles d r, vs. r"' Is 

indirect. however. When photosynthesis and transpiration were measured simultaneously following 

exposure to ureas, triazines, diazines. bipyrldyliums, and simeton, the relative reduction in 

photosynthesis was greater than that c:X transpiration, and commenced sooner (van Oorshcot, 1976). 

This suggested direct herbicidal action at the sites cf photosynthesis. However, for PMA. DSA, 

ioxynll, propanll, nitrofen. and fiuorodifen, photosynthesis and transpiration were inhibited to a similar 

extent and at the same time, implying that stomatal closure restricted uptake of C02 • Similar 

81 
~ 

~'---­·~._..--

variations in the relative reduction d transpirat:-:-'7"-:~ion-vs._p_f:a·w~,,l\1~ were observed for nitrobenzene .Tt.:.t.'S.i ... ~&. ~~"'~j(~lS..•••e• .. 

(McFar1ane et al., 1990). Herbicides also disrupt the water economy d plantS IJy'tWootb'tls!.'l¥OC-t . 

water uptake, as reported for phenoxy herbicides, uraas, triazines, propanl, and dinoseb (van 

Oorschot, 1976). 

Herbicides that inhibit the light reactions d photosynthesis often Inhibit photorespiration as 

well. This has been demonstrated by simiar reductions in net photosynthesis of wheat and com in 

air with normal and low 0 2 concentrations following exposure to dluron (Downton and Tregunna, 

1968), and by the absence of the postllumlnatlon C02 burst In barley exposed to atrazine (lmbamba 

and Moss, 1971). However, when com was exposed to atrazine, C02 evolution in the light was 

greater than in darkness, suggesting little or no suppression of photorespiration (van Oorschot, 1976). 

Other Organics 

Organic chemicals other than herbicides have been shown to Influence plant growth and 

development. Wood et al. (1985) and Wood and Payne {1986) found that pecan leaves sprayed with 

the fungicides propiconazole, benomyt, triphenyltin hydrcxide, and docline, or with pyrethroid, 

carbamate, and organophosphate insecticides showed inhibition of net photosynthesis of up to 20% 

within one day following application, with full recovery usually occurring within several days. 

Hydrocarbon oils universally reduce transpiration and net photosynthesis of crops (Baker, 1970). 

Recovery of both processes Is correlated with the dissipation of foliar-applied oils, and raising the CO2 

concentration of the air reduces inhibition, suggesting that blockage of stomata Is th.e cause of 

inhibition. However, due to the rapidity with which hydrocarbons penetrate leaf tissues, nonstomatal 

effects are likely as well. Hydrocarbon oUs either increase or decrease respiration, with the effect 

differing among ons and between plant species (Baker, 1970). Nitrobenzene at 8 g/L in growth 

solution had varying effects on gas exchange of several species; ranging from no effects on soybean 

to complete photosynthetic repression of green ash (McFarlane et al., 1990). Even at 100 g/L. 

nitrobenzene did not affect soybean gas exchange (Aetcher et al., 1990). 

One hypothesis to account for European forest-tree decline or "Waldsterben· is phytotoxicity 

of anthropogenic and biogenic hydrocarbons. When young Norway spruce were exposed to ,.pinene 

vapors, photosynthesis was Inhibited and chlorophyll degraded (Grosset al., 1988; Wagner et al., 

1989). ,.pinene inhibits the HUI reaction (Pauly, 1981), resLJtlng In uncoupling of oxidative 

phosphorylation (Wagner et al., 1989). Frank and Frank (1985) were able to reproduce the decline 

symptom of needle yellowing In Norway spruce by exposing branches to vapors of two anthropogenic 

VOCs, tri- and tetrachloroethene, in the field. They later showed (1986, 1989) that these compounds 

were absorbed by spruce needles and, in the presence of sunlight, the compounds caused 

photodegradation of photosynthetic pigments. Rippen et al. {1987) have speculated that nitrophenols 

injure forests. 

82 
Ultrastructure and Membrane Function 

Because most organics are llpophUic, It Is not surprising that they often disrupt cell 

membranes. At IOYI concentrations, dissolution d organic moi8CUes in membranes forces apart the 

fatty acid chains d the phospholipids, Increasing membrane permeability. At higher concentrations, 

the bUayer configuration may disintegrate completely. Some organics disrupt cell membranes 

indirectly through the formation d free radicals, Inhibition d lipid synthesis, or changes In the types of 

lipids synthesized. The disruption d membranes probably accounts for many d the lJtrastructural 

changes attributed to organics. 

Herbicides 

Contact foliar damage rbumlng• or "wetness1 by 2,4-0, endothal, and several other 

herbicides is believed to result from physical disruption of cell membranes (Morrod, 1976). Paraquat 

disrupts membranes because Its metabolism produces hydrogen peroxide (Morrod, 1976). Two 

consequences of membrane damage are leakage d cell contents and altered rates of active uptake. 

For example, DNP increases plant cell permeability to water, whereas stomatal closure caused by 

atrazine, 2,4-D, and DNP may be caused by leakage of K .. from guard cells (Morrod, 1976; Shea et 

al., 1983). Ultrastructural changes induced bX herbicides in leaves Include swelling of thylakoids and 

chloroplasts, destruction of phloem and vascular cambium, abnormal differentiation of xylem, bark, 

and root cells, and mitotic abnormalities (Amer and Ali, 1969; Unck, 1976; Moreland and Hilton, 1976}. 

Other organics 

Despite rather low K_ values, bulky hydrocarbons are effective at disrupting membranes. For 

example, heptene-2, with a bent configuration, causes discoloration and death of cotton hypocotyls, 

while straight-chain heptane shows no phytotoxicity (MomJd, 1976). It has also been shown that 

,.pinene alters the relative proportions of lipids and causes disappearance of microsomes in Norway 

spruce needles (Frosch et al., 1989; KJingler and Wagner, 1989}. 

Translocation 

Root ion uptake Is Inhibited by many organics. For example, nltrophenols, 2,4-0, simazine, 

and propanA inhibit phosphate uptake, DNP inhibits Ca. 2-+ uptake, the organophosphorus Insecticide 

Systox inhibits a· uptake, and phenolic acids inhibit phosphate and K + uptake (Barber and Koontz, 

1963; Glass, 1973, 1974; Oertli and Ahmadi, 1975; Kaufman, 1976; Morrod, 1976). As uptake of these 

ions is energy-requiring, the mechanism d reduced uptake may be inhibition of respiration or 

uncoupling of oxidative phosphorylation. 

83 
Herbicides alter the qualitative and quantitative nature d solute transport in plants. The major 

effects are physical obstruction of the phloem caused by unorganized cell dMsion, blockage by 

callose formation, localized xylernjphloem tissue injury or death, stimLAtion or inhibition of metabolic 

sinks, and decreased transpiration (Ashton and Bayer, 1976). The only reference found for a nonherbicide organic was Wedding and Riehl {1958), who reported that petroleum ol inhibited phosphate 

transport from roots to leaves In lemon seedlings. 

Metabolism 

Herbicides 

Herbicides have various modes ci action. Many Interfere with mitochondrial metabolism. 

Chloro- and nitro-phenols, benzlmidazoles, nitriles, and phenoxy acids are uncouplers of oxidative 

phosphorylation O.e., the synthesis of ATP is stopped even though respiration continues) (Beevers, 

1953; Kaufman, 1976; Kirkwood, 1976; Shea et al., 1983). Substituted phenols vary widely in their 

uncoupling ability (Gaur and Beevers, 1959). Other herbicides, Including amides, benzoic acids, 

carbamates, dinitroanilines, halogenated aliphatics, thiocarbamates, triazines. triazoles, and ureas, 

reduce oxidative phosphorylation by either Inhibiting the transfer of energy to intermediates In the 

formation of ATP, or inhibiting the flow of electrons along the electron-transport chain (Kirkwood, 

1976). Glycolysis is Inhibited by halogenated aliphatics and phenoxy acids; the latter also Inhibit the 

pentose-phosphate pathway (Kirkwood, 1976). 

Many herbicides negatively affect photosynthesis. Inhibition of electron transport in 

Photosystems I and II occurs by removal or Inactivation of intermediate electron transport carriers, or 

by the herbicide acting as an electron acceptor in competition with normal acceptors. This Inhibition 

is light-dependent. Uncouplers dissociate electron transport from ATP synthesis 

(photophosphorylation). Some herbicides are both inhibitors and uncouplers. For a comprehensive 

review, see Moreland and Hilton (1976). 

Herbicide effects on Intermediate metabolism are numerous. Many inhibit rates of synthesis 

or degradation of carbohydrates, lipids, proteins, and nucleic acids, or alter the types of these 

compounds fOI'TTlGd (Cherry, 1976; Kirkwood, 1976; Morrod, 1976). However, 2,4-D and 2,4,5-T 

stimulate lipid synthesis (Morrod, 1976). Some herbicides inhibit nitrite reduction, causing nitrite to 

accumulate In plant tissues (Klepper, 1979). 

Other Organics 

Fungicides (e.g., captan) inhibit respiration and photosynthesis at the cellular level !n a similar 

manner to herbicides (Budimir et al., 1976). Evidence exists for inhibition of gylcolysls, aerobic 

84 
respiration, and uncoupling of oxidative phosphorylation by hydrocarbon ols (Baker, 1970). Ravanel 

et al. (1989) reported the uncoupling ability of chlorophenol in maple cell suspensions. 

Plant Resistance 

Plant species differ in their resistance to herbicide effects, and species sensitivity rankings 

differ among herbicides. Other factors Influencing sensitivity are genotype, growth stage, nutritional 

status, physical damage, cuticUar permeablity, temperature, water stress. light Intensity, leaf area, leaf 

morphology, transpiration rate, and rooting depth (Aberg and Stecko, 1976; Muzik, 1976). It Is well 

established that the primary causes of "herbicide selectMt( (the pesticide science term for plant 

sensitivity) are due primarily to three factors: (1) different rates d uptake by the roots and/or shoots; 

(2) different rates of translocation and differences In the organs and tissues to which translocation 

occurs; and (3) different capabilities for, or rates of, metabolic detoxification (Sargent, 1976; Wain and 

Smith, 1976). At the site(s) of herbicide action, species differences In sensitivity often disappear. For 

example, the Hill reaction In Isolated chloroplasts from different plant species is Inhibited to the same 

degree by the same concentrations of simazine and several other herbicides (van Oorschot, 1976). 

Differential resistance of plant species to hydrocarbon ols has also been observed (Baker, 1970}. In 

the case of the Umbelliferae , cell membranes have Inherent high resistance to penetration by oils 

(van Overbeek and Blondeau, 1954). 

It Is worth noting that plants contain a bewUdering variety of organic compounds that are 

toxic to animals. Most of these are believed to be evolved, qualitative or quantitative, defenses 

aga~nst herbivory. Interestingly, many of these are known phytotoxins, Including organic acids and 

phenols (Duke, 19n). This fact clear1y indicates that many plants already possess mechanisms to 

detoxify some priority air pollutants. 

POTENTIAL EFFECTS ON PLANTS 

Ecological Effects 

When photosynthesis and respiration are Inhibited, or ATP synthesis uncoupled from 

metabolism, the whole-plant reaction should be reduced growth and reproduction. Energy allocated 

to alleviating the stress of chronic toxin uptake may also reduce the plants ablity to produce 

defensive chemicals. Inhibition of root water or ion uptake may increase allocation to roots at the 

expense cA shoots. Together, these processes lead to lesser competitive ability of affected species. 

Except near point sources of toxic organics, these effects may occur gradually as small but 

85 
continuous atmospheric Inputs accumulate. Ultimately, shifts In species composition and dominance 

may become apparent at the plant community level. 

Just which species may suffer most Is largely a matter of speculation. For those organics 

that readily cross the root endodermls, rates of translocation, and hence of bioaccum&Jation, may be 

higher for plants with greater transpiration rates O.e., Inherently smaller stomatal resistances). These 

tend to be ear1y successionaJ species with fast growth rates and among late-successional species that 

are shade Intolerant. Conifers have the highest stomatal res!stances d any major plant life form 

(Komer et al., 1979), yet they also have exceptionally thick, waxy cuticles Into which organics may be 

absorbed. The major route of entry, foliage vs. roots, of organics COIJd play a pivotal role In 

determining the relative resistances of different plant taxa. 

Interactions with other Pollutants 

Many organics are weak ack:ts, and their lipophUicity Is greater In the unionized form. 

Penetration of weak-acid herbick:tes through cuticles Increases as solution pH decreases (Fogg, 1948; 

Sargent and Blackman, 1962; Bukovac, 1976). This phenomenon probably accounts for the enhanced 

phytotoxicity of nitrophenols and organic ack:ts at low pH (Simon and Beevers, 1951; Simon et al., 

1952). As a result, there Is a strong possibility that ack:tifled wet deposition Increases the 

phytotoxicity af organics by enhancing uptake (Mullen, 1986). 

Biogenic monoterpenes present an interesting case. Uke anthropogenic VOCs, they 

contribute to the photochemical formation of ozone and PAN (Lurmann et al., 1983; Trainer et al., 

1987; Chameides et al., 1988) that have been shown unequivocally to reduce photosynthesis and 

growth of crop and tree species (Reich, 1987). Wagner et al. (1989) have hypothesized that ozone 

penetrates the mesophyll of conifers, increasing membrane permeability and liberating monoterpenes 

stored in secretory vesicles. In the presence of ozone, monoterpenes promote the oxidation of sulfur 

dioxide to sulfate, possibly causing local enhancement of acid deposition to forests (Stangl et al., 

1988). 

Interactions between different types of gaseous pollutants are well documented (Ormrod, 

1982). Because terrestrial plant communities are receMng chronic Inputs of possibly hundreds of 

different organic compounds, It Is possible that the overall effects In combination are additive, more 

than additive, or less than additive. 

86 
CONCLUSIONS 

Thousands of organic compounds are released Into the atmosphere each year by human 

activities, induding pesticides, aliphatic and aromatic hydrocarbons (both halogenated and 

unhalogenated), polychlorinated biphenyls, and phthalate esters. These are transported long distances 

and reach natural vegetation via wet and dry deposition. Deposition ftuxes and velocities have been 

characterized for only a few organics In a few geographic areas. 

The potential for plant uptake and bioconcentration d organics is high and correlated with 

octanol-water partition coefficients. Greater transpiration rates produce greater root uptake. Foliar 

uptake of vapor-phase organics has received little attention. Translocation patterns differ among 

organic compounds. Metabolic detoxifK'.atlon plays a major role In plant resistance, but has been well 

characterized only for pesticides. 

With few exceptions, the modes cl action cl organics In plants are well-known only for 

herbicides. A variety of effects by herbicides, usually but not always Inhibitory, have been observed 

on plant growth and morphology; photosynthesis, respiration, and transpiration; cell membranes and 

ultrastructure; translocation; and respiratory, photosynthetic, and intermediate metabolism. Many of 

these same effects have been observed with nonherbicide organics. I found no studies on 

population-, community-, or ecosystem-level effects of airborne organics. 

Relevance of Previous Work 

As the analysis of the PHYTOTOX database indicated, there is only one class of organics -

herbicides - for which the physiological effects on plants are relatively well known. This is hardly 

surprising, because their purpose Is to kill plants, especially weeds. Although It would .be desirable to 

know more about nonherbicide organics, existing knowledge of herbicide effects Is helpful In 

predicting the impacts of organics In general on terrestrial vegetation. For example, herbicides are at 

times applied aerially to crops or forests, with consequent downwind drift onto adjacent natural plant 

communities. Some herbicide precursors (e.g., 2,«fichlorophenol, used to make 2,4-0, and maleic 

anhydride, used to synthesize pyrldazines) are on the EPA's Toxic Release Inventory (U.S. EPA, 1989). 

Point releases of these, and ci the finished herbicides themselves, from sites cl manufacture may 

affect natural vegetation. A few herbicides have other human sources besides pesticide manufacture 

and use. For example, DNP, DNOC, and other nitrophenols are produced as by-products cl smogforming reactions. Finally, the observed correlations between plant bioconcentration or modes of 

action and the physicochemical properties ci herbicides permit general predictions to be made 

concerning the phytotoxic effects of nonherbicide organics. 

87 
Nevertheless, the literature emphasis on herbicides poses many obstacles to understanding 

potential phytotoxicity c:K airborne organics. One obstacle is that the conditions under which 

herbicides are applied to plants - high concentrations rNer short time periods - are quite different from 

those encountered by natural vegetation - low concentrations In a repeated or continuous manner. 

Significant differences between acute and chronic exposures to the major airborne pollutants have 

been demonstrated (Lefohn and Runeckles, 1987). Thus, extrapolation from acute agricUtural doses 

to chronic natural vegetation doses may be unwarranted. Another dlfficUty is the extremely high 

concentrations ci organics used In most dose-response experiments. For example, solution 

concentrations d 10 3 to 101 ng/L are frequently encountered In phytotoxicity studies. Compare these 

with the concentrations In Table 1. 

The best use that we may be able to make d the existing literature is to rank the likelihood of 

plant uptake of organics according to the vapor pressures, K_ values, and half-Uves in soU of the 

organics (Ryan et al., 1988; Travis llnd Hattemer-Frey, 1988). 

RECOMMENDATIONS 

The most critical areas in need of further knowledge are: (1) spatial and temporal patterns of 

atmospheric concentrations and wetjdry deposition rates; (2) rates d uptake by roots from soU, and 

of vapors and wet deposition by shoots; (3) translocation patterns; (4) physiological effects from the 

whole-plant to the subcellular level; and (5) modes of action. In addition, much more emphasis needs 

to be placed on nonpesticide organics and on native, nonagricultural species growing in natural 

environments. In all such work, we must use acute or chronic doses that bridge the range of 

exposures actually occurring In natural ecosystems. 

Some soecific suggestions are as follows. 

(1) Upgrade routine air, aerosol, and precipitation monitoring networks (e.g., National Atmospheric 

Deposition Program, National Dry Deposition Network) to measure concentrations d at least 

the most common and highly concentrated organics. 

(2) Characterize dry deposition to various plant communities. Traditional direct measurement 

techniques- flux-gradient and eddy correlation (Hicks, 1986) -are diffiCult due to the very low 

concentrations d most organics. However, new advances In laser technology- FTIR, lidar, 

tunable diode lasers - hold promise for use of the flux-gradient technique in the near future. 

88 
(3) Develop methods to measure short-term uptake ci organics by plant shoots. These would 

com~ement 14C work by distinguishing between cuticWlr and stomatal absorption, and would 

permit scaling-up of leaf~evel uptake to entire canopies In Inferential methods ci estimating 

dry deposition (Hicks, 1986). The technological challenge of generating and measuring low 

but stable concentrations is great, but the potential gain In predictive capability should be 

worth the effort. 

(4) Conduct experiments on dose-response relationships of whole plants growing in native soils, 

Including growth, allocation, morphology, water relations, carbon assimlation, and nutrition. 

Growth chamber, continuously stirred tank reactor (CSTR), and open-top chamber studies, in 

order, would provide increasing realism In terms of environmental conditions. 

(5) Determine the cellular ;subcellular modes of action of various classes of organics. Previous 

work with herbicides suggests that virtually every translocation and metabolic process Is 

worthy of investigation. 

(6) Measure dose-response relationships for species representative ci other major life forms of 

terrestrial ~nts besides crops: annual and biennial weeds, perennial herbs, ferns, evergreen 

and deciduous shrubs, evergreen and deciduous hardwoods, and conifers. 

(7) Develop sensitivity rankings based on absorbed rather than external dose. This approach has 

yielded better predictive capability for major gaseous air pollutants (Reich, 1987). Because 

organics vary so widely in their lipophRicity, even within similar structural classes, relationships 

between external dose and response for one organic are unlikely to ap~y to other organics. 

89 
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